centlib.ajums.ac.ircentlib.ajums.ac.ir/multiMediaFile/2223281-4-1.pdfRelated Titles Parlar, H.,...

Organic Indoor Air Pollutants

Occurrence, Measurement, Evaluation

Edited byTunga Salthammer and Erik Uhde

Second, Completely Revised Edition


Edited by

Tunga Salthammer

and Erik Uhde

Related Titles

Parlar, H., Greim, H. (eds.)

The MAK-Collection for Occupational Health and SafetyPart III: Air Monitoring Methods, Volume 11

Series: The MAK-Collection for Occupational Health and Safety. Part III: Air Monitoring

Methods (DFG) (Volume 11)

2009

ISBN: 978-3-527-31959-6

Parlar, H., Greim, H. (eds.)

The MAK-Collection for Occupational Health and SafetyPart III: Air Monitoring Methods, Volume 10

Series: The MAK-Collection for Occupational Health and Safety. Part III: Air Monitoring

Methods (DFG) (Volume 10)

2007

ISBN: 978-3-527-31601-4

Bester, K.

Personal Care Compounds in the Environment

2007

ISBN: 978-3-527-31567-3


Occurrence, Measurement, Evaluation

Edited byTunga Salthammer and Erik Uhde

Second, Completely Revised Edition

The Editors

Prof. Dr. Tunga SalthammerFraunhoferWilhelm-Klauditz-Institut (WKI)Material Analysis and Indoor ChemistryBienroder Weg 54 E38108 BraunschweigGermany

Dr. Erik UhdeFraunhoferWilhelm-Klauditz-Institut (WKI)Material Analysis and Indoor ChemistryBienroder Weg 54 E38108 BraunschweigGermany

All books published by Wiley-VCH are carefully produced. Nevertheless, authors, editors, and publisher do not warrant the information contained in these books, including this book, to be free of errors. Readers are advised to keep in mind that statements, data, illustrations, procedural details or other items may inadvertently be inaccurate.

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Bibliographic information published by the Deutsche NationalbibliothekDie Deutsche Nationalbibliothek lists this publication in the Deutsche Nationalbibliografi e; detailed bibliographic data are available on the Internet at <http://dnb.d-nb.de>.

© 2009 WILEY-VCH Verlag GmbH & Co. KGaA, Weinheim

All rights reserved (including those of translation into other languages). No part of this book may be reproduced in any form – by photoprinting, microfi lm, or any other means – nor transmitted or translated into a machine language without written permission from the publishers. Registered names, trademarks, etc. used in this book, even when not specifi cally marked as such, are not to be considered unprotected by law.

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Printed in the Federal Republic of Germany Printed on acid-free paper

ISBN: 978-3-527-31267-2

http://dnb.d-nb.de

V

During my lectures on ‘ indoor air quality ’ with architectural students, I often ask them how much, in their opinion, air weighs. The most common answer I get is that 1 m 3 of air must weigh just 1 g or less. They believe that air is very light. However, the weight of 1 m 3 of air is 1.2 kg. Our daily breathing rate is 15 ∼ 20 m 3 of air – approximately 0.3 m 3 per 1 kg of body weight. Thus, we inhale and exhale approximately 20 kg of substances every day. The mass of inhaled air is much more than that of drinking water and food.

Materials made from organic compounds contribute to improvement of the quality of life; on the other hand, organic chemical pollutants emitted from materi-als and appliances can adversely affect human health. People in developed coun-tries spend more than 90% of their time indoors. In the light of this fact, the cleanliness of occupied spaces such as buildings, houses, and transportation systems becomes very important. In contemporary society it can be assumed that the quality of building products, houses and equipment is relatively poor. More-over, people often suffer from pollutants caused by activities like cooking, cleaning and heating.

The conservation of energy is strongly recommended from the viewpoint of saving the global environment. An air - conditioning system is often installed to obtain thermal comfort indoors; as a result, there is a marked increase in energy consumption for cooling, heating, and ventilation. With regard to buildings and their environments, the increase in life - cycle CO 2 emissions has often been dis-cussed in recent years. Approximately 40% of CO 2 is emitted from the building sector, including housing. Therefore, reducing the emission of CO 2 from the building sector is imperative to prevent global warming.

Air - tightness and insulation are effective measures for energy conservation. Reduction in ventilation in air - conditioned spaces is often considered to be one of the most effective methods to conserve energy. However, as a result of lower air exchange rates, the indoor air concentrations of pollutants, such as volatile organic compounds (VOCs) and semi - volatile organic compounds (SVOCs) emitted from building materials and other sources, increase. This often leads to building - related symptoms if the dwell time in a polluted indoor environment is high.

During the 1980s in Europe and North America and the 1990s in Japan, indoor air pollution by formaldehyde was identifi ed and suitable countermeasures were

Organic Indoor Air Pollutants. 2nd Edition. Edited by Tunga Salthammer and Erik UhdeCopyright © 2009 WILEY-VCH Verlag GmbH & Co. KGaA, WeinheimISBN: 978-3-527-31267-2

Foreword

VI Foreword

taken. Formaldehyde is a single chemical compound, and since we are already aware of the sources of emission, it is comparatively easy to control it. However, since VOCs and SVOCs consist of many substances, it is diffi cult to control their emissions effectively. VOCs and SVOCs are also emitted from natural materials. Moreover, a proper health risk assessment of VOC mixtures has not yet been established.

Indoor air quality is an important determinant of health and well - being. To maintain better indoor air quality, we have to understand the mechanism of indoor air pollution. For this purpose, the measurement of indoor air concentration and use of chemical analysis methods are essential. To estimate indoor air concentra-tion, we have to know the emission and ventilation rates. Emission takes place not only from building products but also from automobile parts, electric appliances, offi ce equipment such as printers, household consumer products, and even printed materials like newspapers. This book serves as a useful guide for chemists, architects, mechanical engineers, constructors, and manufacturers of electronic products. It emphasizes a holistic and multidisciplinary approach toward the indoor environment.

This book reminds us that a healthy indoor environment is essential, and pro-vides scientifi c evidence and countermeasures for the future.

Shin - ichi Tanabe, Prof., Ph.D. Department of Architecture,Waseda University, Tokyo, Japan

VII

Contents

Foreword V Preface to the Second Edition XVII List of Contributors XIX List of Symbols and Abbreviations XXIII

Part One Measuring Organic Indoor Pollutants

1 Application of Solid Sorbents for the Sampling of Volatile Organic Compounds in Indoor Air 3

Erik Uhde1.1 Introduction 31.2 Solid Sorbents – A Brief Overview 41.3 Active or Passive Sampling 71.4 Thermal Desorption or Solvent Extraction 81.5 Sampler Design 81.6 Breakthrough Volumes 111.7 Safe Sampling Volume 111.8 Artifacts and Interferences 121.8.1 Water Affi nity – A Chromatographic Problem 121.8.2 Sorbent Degradation Products and Sorbent Background 131.8.3 Target Compound Degradation and Artifact Formation 151.9 Conclusions 16

2 Sampling and Analysis of SVOCs and POMs in Indoor Air 19 Per Axel Clausen, Vivi Kofoed-Sørensen2.1 Introduction 192.2 Defi nitions and Properties of SVOCs and POMs 192.2.1 Gas/Particle Partitioning in Indoor Air 202.2.2 Surface Adsorption 212.2.3 Health Related Properties 222.3 Compounds and Matrices in the Indoor Environment 222.4 Sampling, Transport and Storage of SVOC/POM Samples 232.4.1 Preparation of Sampling and Analysis Equipment 23


VIII Contents

2.4.1.1 Background Contamination and Loss of Target Compounds 232.4.1.2 Cleaning of Filters 242.4.1.3 Cleaning of Sorbents 242.4.1.4 Cleaning of Glassware and Other Equipment 242.4.2 Sampling SVOCs/POMs in Air 252.4.2.1 Filter/Sorbent Sampling 252.4.2.2 Determination of the Gas/Particle Partitioning: Denuder Sampling 262.4.2.3 Artifact Formation Caused by Reactive Gases in Indoor Air 262.4.2.4 Air Sampling Pumps 272.4.3 SVOCs/POMs in Surface Dust 272.4.3.1 Filter Sampling with Vacuum Cleaner 272.4.3.2 Specially Designed Dust Sampler 282.4.4 SVOCs/POMs in Building Materials and Consumer Products 282.4.4.1 Indoor Material Samples Containing SVOCs/POMs 282.4.4.2 Testing Emission of SVOCs from Indoor Materials in Chambers 282.5 Preparation of SVOC/POM Samples for Analysis 302.5.1 Extraction of SVOCs/POMs from Samples 302.5.1.1 Cleaning of Extraction Equipment 312.5.2 Concentrating Extracts of SVOC/POM Samples 322.6 Analysis of SVOCs/POMs 322.6.1 Gas Chromatography (GC) 322.6.1.1 On-Column Injection (OC) 342.6.1.2 Large Volume Injection (LVI) 342.6.1.3 Thermal Desorption (TD) 342.6.1.4 ‘Cold Spots’ and Other Adsorption Problems 352.6.1.5 Flame Ionization Detection (FID) 352.6.1.6 Mass Spectrometric Detection (MS) 352.6.2 High Performance Liquid Chromatography (HPLC) 362.6.2.1 HPLC with Fluorescence Detection (HPLC-FD) 362.6.2.2 HPLC with Mass Spectrometric Detection (LC-MS) 362.6.3 Analysis Sequences 362.7 Quality Assurance and Control 372.7.1 Method Validation 372.7.1.1 Calibration Curves 392.7.1.2 Limit of Detection (LD) and Limit of Quantifi cation (LQ) 392.7.2 Controls and Control Charts 412.7.3 Documentation 41 References 42

3 Application of Diffusive Samplers 47 Derrick Crump3.1 Introduction 473.2 Principles of Diffusive Sampling 483.3 Selection of Appropriate Methods 503.4 Performance of Diffusive Samplers for the Measurement of VOCs in

Indoor Air 50

Contents IX

3.5 Studies of VOCs in Indoor Air Using Diffusive Samplers 533.6 Other Applications of Diffusive Samplers 593.7 Conclusion 59 References 60

4 Real-Time Monitoring of Indoor Organic Compounds 65 Yinping Zhang, Jinhan Mo4.1 Introduction 654.2 Proton Transfer Reaction – Mass Spectrometer (PTR–MS) 664.2.1 Detection Principles 664.2.2 Measuring Method 684.2.3 Accuracy, Linearity, Limits of Detection and Precision 694.2.4 Applications of PTR–MS 724.3 Photo-acoustic Spectroscopy 734.3.1 Detection Principles 734.3.2 Measuring System and Method 744.3.2.1 Discrete Sampling: Nondispersive PAS 744.3.2.2 Discrete Sampling: FTIR/PAS 764.3.2.3 Continuous Flow-PAS 764.3.3 Selectivity, Sensitivity and Accuracy 774.3.4 Applications of PAS 784.4 Flame Ionization Detection 784.4.1 Detection Principle 794.4.2 Measuring System and Method 794.4.3 Selectivity and Sensitivity 804.4.4 Applications of FID 804.5 Photo-ionization Detection 804.5.1 Detection Principles 814.5.2 Selectivity and Sensitivity 814.5.3 Applications of PID 824.6 Metal Oxide Sensors 834.6.1 Measuring Principle 834.6.2 Selectivity and Sensitivity 864.7 Air Sampling and Data Recording 874.8 Examples of Investigations Using Real-Time Monitoring 874.8.1 Laboratory Investigations of VOC Emissions from Building

Materials 874.8.1.1 Experimental Principle 884.8.1.2 Experimental System 884.8.2 Organic Compounds in Outdoor Air 904.8.3 The Effect of Photocatalytic Oxidation on VOC Removal 914.8.3.1 Detection of Harmful By-Product During the Removal of Toluene by

PCO 924.8.3.2 Evaluating the Formaldehyde Removal Performance of PCO

Reactors 94

X Contents

4.8.4 Products of Ozone-Initiated Chemistry in a Simulated Aircraft Environment 94

4.9 Concluding Remarks 96 Acknowledgments 97 References 97

5 Environmental Test Chambers and Cells 101 Tunga Salthammer5.1 Introduction 1015.2 Characteristics of Chambers and Cells 1025.3 Sink Effects 1055.4 Calculation of Emission Rates 1065.5 Kinetics and Mass Transfer 1085.6 Application of Test Chambers and Cells 1095.7 Final Remarks 112 References 113

Part Two Investigation Concepts and Quality Guidelines

6 Standardized Methods for Testing Emissions of Organic Vapors from Building Products to Indoor Air 119

Elizabeth Woolfenden6.1 Introduction: The Need for Standardization 1196.2 Materials Emissions Testing: A Challenge for Method

Standardization 1206.2.1 The Range of Products and Materials Requiring Emissions

Testing 1216.2.2 The Range of Potential Target Compounds 1216.2.3 Method Variability or Uncertainty 1306.2.4 Nonuniformity of Test Methods 1306.3 Regulations, Standard Methods and Test/Certifi cation

Protocols 1316.4 Emissions Test Methods for VOCs: An Overview of Basic

Principles 1336.4.1 Standard test Methods for Formal Evaluation and Certifi cation of

Emissions 1336.4.2 Secondary or ‘Screening’ Methods for Materials Emissions 1346.5 The Total-VOC Debate 1376.6 Standard Methods and Protocols for Emissions Testing: Current

Status 1386.6.1 Typical Conditions for Emissions Testing Using

Chambers/Cells 1386.6.2 Standard Methods: What Can Go Wrong? 1396.6.2.1 Effect of the Emission Mechanism 139

Contents XI

6.6.2.2 Collection and Transport of Samples Plus Homogeneity Issues 1406.6.2.3 Potential Variables Associated with Testing Materials Using Emissions

Chambers/Cells: Edge Effects, Sample Orientation and Sample Storage Between Tests 140

6.6.2.4 Sink Effects 1416.6.2.5 Target Analytes and System Calibration 1416.6.2.6 Chromatographic Integration and Summation Limit Levels 1426.7 Confi dence Limits for Emissions Test Data for Individual VOCs 1436.8 Concluding Remarks 143 Acknowledgments 144 References 144

7 Standard Test Methods for the Determination of VOCs and SVOCs in Automobile Interiors 147

Michael Wensing7.1 Introduction 1477.2 Conditioning of the Automobile Interior 1497.3 Measurement Procedure 1517.3.1 Quantitative Determination 1527.3.2 Semi-Quantitative Determination of VOCs (TVOC) 1547.3.3 Qualitative Determination of VOCs (Identifi cation) 1547.3.4 Identifi cation of SVOCs (Fogging Precipitate) 1557.3.5 Measurement of the Sum of Organic Substances (ΣVOC) 1557.4 Quantitative and Qualitative Results from Brand New Cars 1567.5 Emissions of Organophosphate Esters inside Automobiles 1597.6 Conclusion 161 References 161

8 Material and Indoor Odors and Odorants 165 Florian Mayer, Klaus Breuer, Klaus Sedlbauer8.1 Introduction 1658.2 Odor Evaluation 1678.2.1 Indoor Environments 1678.2.2 Materials 1688.2.3 Panels and Scales 1688.3 Odor Analysis – Odorant Identifi cation 1728.3.1 Methods 1728.3.1.1 Sampling of Volatiles and Odorants from Indoor Environments 1748.3.1.2 Sampling of Volatiles and Isolation of Odorants from Materials 1758.3.1.3 Identifi cation 1758.3.2 Examples 1768.3.2.1 Cleaning Products, Detergents, Air Fresheners 1768.3.2.2 Carpets 1768.3.2.3 Adhesives 177

XII Contents

8.3.2.4 Rubber Materials Used for Sealings, Floorings, Insulations 1778.3.2.5 Wood 1778.3.2.6 Wood-Based Flooring Materials 1788.3.2.7 Linoleum 1788.3.2.8 Gypsum-Based Products 1798.3.2.9 Plastics 1798.3.2.10 Electronic Devices 1808.3.3 Odorants and Odor Thresholds 1808.3.4 Application of the Combination of Odor Evaluation and Odor Analysis

for Product Optimization 1828.4 Conclusion and Outlook 183 References 184

9 Evaluation of Indoor Air Contamination by Means of Reference and Guide Values: The German Approach 189

Birger Heinzow, Helmut Sagunski9.1 Introduction 1899.2 Defi nition of Terms 1909.2.1 Indoor Environment 1909.2.2 Utilization Cycle 1909.2.3 Volatile Organic Compounds (VOCs) 1919.3 Values for Evaluating the Indoor Air Quality 1919.3.1 Toxicologically Based Values 1919.3.2 Statistically Defi ned Values 1929.4 Evaluation of Indoor Air Quality with the Aid of Guide Values 1929.4.1 Requirements Relating to Guide Values for Indoor Air 1929.4.1.1 Health Reference 1929.4.1.2 Legal Reference 1949.4.2 Basic Scheme for Deriving Guide Values for Indoor Air 1949.4.3 Application of the Guide Values in Risk Management 1969.4.4 Recommendation 1979.4.5 Guide Values by the Ad-hoc WG Not Based on RW I and RW II 1979.5 Health Evaluation with the Aid of the TVOC Concept 1989.5.1 Recommendation Relating to the Application of TVOC Values 1989.5.2 Time Curve of Higher TVOC Concentrations 2039.6 Evaluation of Indoor Air Quality with the Aid of Reference

Values 2039.6.1 The Current State of Indoor Air Reference Values 2049.6.2 Recommendations 2049.7 Application of Measured Values in Order to Evaluate Indoor Air

Quality 2069.8 Evaluation of Substances Without Reference Values From the IRK/

AOLG Ad-hoc Working Group 207 Acknowledgment 208 References 209

Contents XIII

Part Three Field Studies

10 Effect of Ventilation on VOCs in Indoor Air 215 Kwok Wai Tham, S. Chandra Sekhar, Mohamed Sultan Zuraimi10.1 Introduction 21510.1.1 Building and Ventilation Characteristics of Offi ce Buildings in a

Tropical Climate 21610.2 VOC Concentration Levels in Eight Singapore Buildings 21610.2.1 Concentrations 21710.2.2 Health Effects Caused by VOCs in Singapore Buildings 22110.2.3 Possible Sources 22110.3 Apportionment of VOCs Source Strengths in Five

Buildings 22110.3.1 Area-Specifi c Emission Rates of VOCs 22110.3.2 Source Apportionment of VOC Sources 22510.4 Effects of Typical Ventilation Operations on TVOC Levels 22710.5 Effect of Purging on Indoor TVOC Levels 23010.5.1 Purging System 23010.5.2 Building Characteristics 23110.5.3 Purging Measurements 23310.6 Summary 236 References 237

11 Occurrence of Semi-Volatile Organic Compounds in the Indoor Environment 239

Werner Butte11.1 Introduction 23911.2 Concentrations of SVOCs in Indoor Air and House Dust 24011.2.1 Phenols and Their Derivatives (Other than Biocides) 24011.2.2 Biocides 24111.2.3 Musk Compounds 24211.2.4 Organophosphates 24311.2.5 Organotin Compounds 24611.2.6 Perfl uorinated Compounds 24611.2.7 Phthalates 24811.2.8 Polybrominated Diphenyl Ethers 25311.2.9 Polychlorinated Biphenyls 25311.2.10 Polychlorinated Dioxins and Furans 25611.2.11 Polycyclic Aromatic Hydrocarbons 25711.3 Sources for SVOCs Indoors 26011.4 The Indoor Environment: A Source for Exposure? 26111.4.1 Indoor Air and House Dust: Associations to Human

Biomonitoring 26111.4.2 Indoor Biocides: A Reason for Health Impairments? 26211.4.3 Reference and Guideline Values 263

XIV Contents

11.5 Summary 264 References 265

12 Indoor Pollutants in the Museum Environment 273 Alexandra Schieweck, Tunga Salthammer, Simon F. Watts12.1 The Museum Environment: An Introduction 27312.2 Climatic Conditions 27612.2.1 Humidity 27712.2.2 Temperature 27812.3 Inorganic Atmospheric Compounds 27812.4 Formaldehyde, Organic Acids (Formic Acid, Acetic Acid) 28112.5 Volatile Organic Compounds (VOCs) 28412.6 Semi-volatile Organic Compounds (SVOCs) 28712.7 Occurrence of Biocides in the Museum Environment 28812.8 The Role of People 29112.9 Risk Assessment and Preservation Strategies 29212.9.1 Recommendations and Guidelines 29312.10 Conclusion 293 References 296

13 Indoor Organic Chemistry 301 Glenn Morrison13.1 Introduction 30113.2 Relevance of Chemistry Using Indoor Air Models 30213.3 Homogeneous Chemistry 30313.3.1 Gas-Phase Organic Oxidation Chemistry: Ozone 30313.3.2 Gas-Phase Organic Oxidation Chemistry: Hydroxyl Radical 30813.3.3 Gas-Phase Organic Oxidation Chemistry: Nitrate Radical 30913.3.4 Condensed-Phase Chemistry: Oxidation 31013.3.5 Condensed-Phase Chemistry: Hydrolysis 31113.4 Heterogeneous Chemistry 31313.4.1 Heterogeneous Chemistry: Ozone and Fresh Indoor Surfaces 31313.4.2 Heterogeneous Chemistry: Ozone and Soiled Surfaces 31613.4.3 Heterogeneous Chemistry: Acid–Base 31813.5 Concluding Remarks 319 References 320

14 Human Responses to Organic Air Pollutants 327 Lars Mølhave14.1 Introduction 32714.2 VOC Exposures Indoors 32914.2.1 Health Effects of Indoor Air Pollution 33014.2.2 Indicators of Indoor Air Quality and Health 33214.2.3 Classes of Indoor Air Pollutants 33414.2.4 The TVOC Indicator 336

Contents XV

14.3 Summary of Experimental Evidence of Health Effects of VOC Exposure 337

14.3.1 Symptoms Relevant to VOCs 33714.3.2 Effect of Exposure Types 34214.4 Conclusions 342 References 343

Part Four Emission Studies

15 Volatile Organic Ingredients in Household and Consumer Products 349

Godwin A. Ayoko15.1 Introduction 34915.2 Literature Survey 35015.3 Product Classes 35115.3.1 Newspaper and Journals 35115.3.2 Insecticides 35615.3.3 Air Fresheners and Deodorizers 35715.3.4 Cleaning Agents 35815.3.5 Polishes 35915.3.6 Products for Personal Hygiene and Cosmetics 36115.3.7 Incenses 36315.3.8 Perfumes and Fragrances 36515.3.9 Cooking and Cooking Related Products 36615.3.10 Miscellaneous Products and Studies 36615.4 Conclusion 368 References 368

16 Building Products as Sources of Indoor Organic Pollutants 373 Stephen K. Brown16.1 Introduction 37316.2 Organic Pollutants Emitted from Major Building Products 37316.2.1 Building Products 37316.2.2 Organic Pollutants 37416.2.3 VOC Emissions Levels Over Time 37516.2.4 VOC Emission Limits/Labels 37616.2.5 TVOC Emissions from Building Materials 37716.3 Interior Paints 37716.3.1 Water-Based Paints 37916.3.2 Solvent-Based Coatings 38316.3.3 ‘Natural’ Paints 38616.3.4 Low-VOC/VOC-Free Paints 38716.4 Floor Covering Systems 38816.4.1 Adhesives 388

XVI Contents

16.4.2 Carpets and Underlays 38916.4.3 Plastic Floorcoverings 39216.5 Concrete and Plaster Products 39316.6 Wood-Based Panels 39416.7 Natural Wood 39616.8 Ovens and Heaters 39716.9 Concluding Remarks 399 References 400

17 Emission of VOCs and SVOCs from Electronic Devices and Offi ce Equipment 405

Tobias Schripp Michael Wensing17.1 Introduction 40517.2 Test Procedures 40817.3 VOC and SVOC Emissions from Various Devices 41417.3.1 Printers and Copiers 41417.3.2 Personal Computers 41917.3.3 Television Sets and Computer Monitors 42117.4 Ultra-Fine Particle Emission from Offi ce Devices 425 References 427

Index 431

XVII

Preface to the Second Edition

The fi rst edition of this book went to print in 1999, the year that the 8th Confer-ence on Indoor Air Quality and Climate was held in Edinburgh, Scotland. The papers read at this last major indoor air conference of the fi nal years of the 20th century dealt once again with the central concerns of indoor air research in the 1990s, most of which are also found in the various chapters of the fi rst edition. As regards determination of volatile organic compounds (VOCs), the defi nition of the sum parameter TVOC by an European Union work group and the standardiza-tion of emissions test chambers and cells by a CEN committee may be regarded as milestones. With the introduction of the TVOC value, GC/MS thermal desorp-tion also fi nally established itself as a standard method of analysis. Furthermore, during this period important fundamentals were laid down for the derivation of indoor air guide values and for product labeling.

Over the last ten years, indoor air research has experienced a signifi cant trans-formation and this has made substantial revisions necessary for the 2nd edition of this book. A number of chapters dealing with the topics of solid sorbents, passive sampling, automobile interiors and household products could, after updat-ing, also be included. However, many sampling techniques and analytical methods today form part of the tools routinely used in indoor air research and are for this reason no longer treated in such detail in this new edition. On the other hand, real - time methods, sensory testing and SVOC analysis have gained in importance and have now been given their own chapters. At the present moment probably the highest level of research activity is to be found in the fi eld of indoor chemistry. Although chemical reactions in indoor air were recognized more than 15 years ago as the source of air - polluting substances, systematic investigations have not been possible until relatively recently when the necessary measuring technology became available. A new chapter provides an overview of the state of development in this fi eld. Two contributions are concerned with the effects on health of VOCs and SVOCs as regards exposure and the identifi cation of guide values. Other chapters are devoted to further topics of current interest in the fi eld of indoor air research such as ventilation concepts, museums and archives and also emissions from electronic devices.


XVIII Preface to the Second Edition

Despite all of these changes the book still has the same demand: to provide the reader with a clear introduction to an important area of indoor air research. We should like to express our thanks to all colleagues and friends who as authors submitted state - of - the - art contributions despite their daily workloads and other commitments. We also thank Mrs Lesley Belfi t of the WILEY - VCH Verlag for her support and patience and Frau Susanne Beerstecher for organizational work.

Braunschweig, May 2009 Tunga Salthammer Erik Uhde

XIX

List of Contributors

Godwin A. Ayoko Queensland University of Technology School of Physical and Chemical Sciences International Laboratory for Air Quality and Health GPO Box 2434 Brisbane 4001, QLD Australia

Klaus Breuer Fraunhofer Institute for Building Physics (IBP) Fraunhoferstr. 10 83626 Valley Germany

Stephen K. Brown CSIRO Sustainable Ecosystems PO Box 56 Highett Victoria 3190 Australia

Werner Butte Universit ä t Oldenburg Fakult ä t V, Institut f ü r reine und angewandte Chemie Carl - von - Ossietzky - Str. 9 - 11 26129 Oldenburg Germany


Per Axel Clausen National Research Centre for the Working Environment Lers ø Parkall é 105 2100 K ø benhavn Ø Denmark

Derrick Crump Cranfi eld University Institute of Environment and Health Cranfi eld Health Cranfi eld Beds, MK43 0AL UK

Birger Heinzow Landesamt f ü r soziale Dienste Dezernat Umweltbezogener Gesundheitsschutz Brunswikerstr. 4 24105 Kiel Germany

University of Notre Dame School of Medicine Sydney Campus Australia

Vivi Kofoed - S ø rensen National Research Centre for the Working Environment Lers ø Parkall é 105 2100 K ø benhavn Ø Denmark

XX List of Contributors

Florian Mayer Fraunhofer Institute for Building Physics (IBP) Fraunhoferstr. 10 83626 Valley Germany

Jinhan Mo Tsinghua University College of Architecture Department of Building Science & Technology Beijing 100084 China

Lars M ø lhave Aarhus University Department of Environmental and Occupational Medicine Bartholins All é 2 BLd 1260 8000 Aaarhus Denmark

Glenn C. Morrison Missouri University of Science & Technology Civil, Architectural and Environ-mental Engineering 221 Butler - Carlton Hall 1401 N. Pine St. Rolla, MO 65409 USA

Helmut Sagunski Beh ö rde f ü r Soziales, Familie, Gesundheit und Verbraucherschutz Billstra ß e 80 20539 Hamburg Germany

Tunga Salthammer Fraunhofer Wilhelm - Klauditz - Institute (WKI) Material Analysis and Indoor Chemistry Bienroder Weg 54 E 38108 Braunschweig Germany

Alexandra Schieweck Fraunhofer Wilhelm - Klauditz - Institute (WKI) Material Analysis and Indoor Chemistry Bienroder Weg 54 E 38108 Braunschweig Germany

Tobias Schripp Fraunhofer Wilhelm - Klauditz - Institute (WKI) Material Analysis and Indoor Chemistry Bienroder Weg 54 E 38108 Braunschweig Germany

Klaus Sedlbauer Fraunhofer Institute for Building Physics (IBP) Fraunhofer Stra ß e 10 83626 Valley Germany

Chandra Sekhar National University of Singapore School of Design and Environment Department of Building 4 Architecture Drive Singapore 117566 Singapore

List of Contributors XXI

Kwok Wai Tham National University of Singapore School of Design and Environment Department of Building 4 Architecture Drive Singapore 117566 Singapore

Erik Uhde Fraunhofer Wilhelm - Klauditz - Institute (WKI) Material Analysis and Indoor Chemistry Bienroder Weg 54 E 38108 Braunschweig Germany

Simon F. Watts Victoria University Wellington Department of Geographical, Earth and Environmental Sciences 2 Kelburn Parade Kelburn Wellington 6012 New Zealand

Michael Wensing Fraunhofer Wilhelm - Klauditz - Institute (WKI) Material Analysis and Indoor Chemistry Bienroder Weg 54 E 38108 Braunschweig Germany

Elizabeth Woolfenden Markes International Ltd. Gwaun Elai Campus Llantrisant RCT CF72 8XL UK

Yinping Zhang Tsinghua University College of Architecture Department of Building Science & Technology Beijing 100084 China

Mohamed Sultan Zuraimi National University of Singapore School of Design and Environment Department of Building 4 Architecture Drive Singapore 117566 Singapore

XXIII

List of Symbols and Abbreviations

A sample surface ACH air exchange ADI acceptable daily intake AM arithmetic mean APM airborne particulate matter APS aerodynamic particle sizer AQG air quality guidelines ASHRAE American Society of Heating, Refrigerating and Air -

Conditioning Engineers ASTM American Society for Testing and Materials BaP benzo[a]pyrene BHT 2,6 - di - tert - butyl - 4 - methyl - phenol BREC building related environmental complaints BRI building related illness BRS building related symptoms BTV breakthrough volume C or C (t) concentration C 0 initial concentration C S vapor pressure [mg/m 3 ] CAPs concentrated air particles CEN European Committee for Standardization CFU colony forming unit CIB National Council for Building Research, Studies and

Documentation CMD count median diameter CPC condensation particle counter d distance δ boundary layer thickness D molecular diffusity (diffusion coeffi cient) D p particle diameter DBP di - n - butylphthalate DEHP di - (2 - ethylhexyl)phthalate DIBP di - isobutylphthalate


XXIV List of Symbols and Abbreviations

DINP di - isononylphthalate DMA differential mobility analyzer DMPS differential mobility particle sizer DNPH dinitrophenylhydrazine DOP di - octylphthalate EC electrostatic classifi er ECA European Collaborative Action ECD electron capture detector EDS energy dispersive spectrometer EDXRF energy dispersed X - ray fl uorescence ELISA enzyme linked immunosorbent assay ELPI electrical low pressure impactor EM electron microscopy EPA Environmental Protection Agency EPXMA electron probe X - ray microanalysis ETS environmental tobacco smoke FAAS fl ame atomic absorption spectrometry FID fl ame ionization detector FLEC fi eld and laboratory emission cell FT - IR Fourier transform infrared spectroscopy GC gas chromatography GFAAS graphite furnace atomic absorption spectrometry GM geometric mean GSD geometric standard deviation h height HDM house dust mite HPLC high performance liquid chromatography HVAC heating ventilating air conditioning system IAP indoor air pollution IAQ indoor air quality IC ion chromatography ICP - AES inductively coupled plasma – atomic emission spectrometry ICP - MS inductively coupled plasma – mass spectrometry INAA instrumental neutron activation analysis ISIAQ International Society of Indoor Air Quality and Climate ISO International Organization for Standardization I/O indoor/outdoor k 1 source strength i 2 air exchange (modeling) k i (i > 2) rate constant L loading factor [m 2 /m 3 ] LOAEL lowest observed adverse effect level LOEL lowest observed effect level LOI loss on ignition LR leak rate (test chamber) [h − 1 ]

List of Symbols and Abbreviations XXV

m mass M mass in source M 0 initial mass in source MCS multiple chemical sensitivity MD median MS mass spectrometry MVOC microbiological originated volatile organic compounds N (or n ) air exchange rate [h − 1 ] N p particle concentration NDIR non - dispersive infrared NOAEL no observed adverse effect level NOEL no observed effect level OEL occupational exposure limit OPC optical particle counter OSHA Occupational Safety and Health Administration OT odor threshold [mg/m 3 ] P percentile PAD photo - acoustic detector PAH polycyclic aromatic hydrocarbon PAN peroxyacetyl nitrate PAS photoelectric aerosol sensor (or photoacoustic spectroscopy) PCB polychlorinated biphenyl PCDD polychlorinated dibenzo - p - dioxin PCDF polychlorinated dibenzofuran PCP pentachlorophenol PCR polymerase chain reaction PID photo - ionization detector PIXE particle induced X - ray emission PM particulate matter PM 2.5 suspended particulate matter ( < 2.5 μ m) PM 10 suspended particulate matter ( < 10 μ m) POM particulate organic matter PPN peroxypropionyl nitrate PUF polyurethane foam q area specifi c fl ow rate (N/L) Q air fl ow rate [ml/min, l/h] QSAR quantitative structure - activity relationship r correlation coeffi cient RH relative humidity RI retention index RPM respirable particulate mass RSD relative standard deviation RSP respirable suspended particulate RT retention time SBS sick building syndrome

XXVI List of Symbols and Abbreviations

SD standard deviation or solvent desorption SEM scanning electron microscopy SER A or SER A (t) area specifi c emission rate SER u or SER u (t) unit specifi c emission rate SER V or SER V (t) volume specifi c emission rate SER 1 or SER 1 (t) length specifi c emission rate SIM single ion modus SMPS scanning mobility particle sizer SOA secondary organic aerosol Σ VOC sum of volatile organic compounds SVOC semi - volatile organic compound t time T temperature TD thermal desorption TE thermal extraction TEM transmission electron microscopy TEOM tapered element oscillating microbalance TIC total ion chromatogram TLV threshold limit value TNF α tumour necrosis factor α TSP total suspended particles TVOC total volatile organic compounds (use only as defi ned by ECA

report no 19 !) μ m micrometer μ g microgram v air velocity V volume VDI Verein Deutscher Ingenieure VOC volatile organic compound VVOC very volatile organic compound WAGM weighted average geometric mean WDXRF wavelength dispersed X - ray fl uorescence WHO World Health Organisation XRF X - ray fl uorescence

Measuring Organic Indoor Pollutants


Part One

Application of Solid Sorbents for the Sampling of Volatile Organic Compounds in Indoor Air Erik Uhde

1

1.1 Introduction

The use of building materials, furniture, carpets and household products produces an almost ubiquitous level of volatile organic compound s ( VOC s) in indoor air. Several hundred different compounds have been identifi ed in the indoor environ-ment. Since most air pollutants occur in low concentrations of 1 – 1000 μ g/m 3 , highly sensitive detection methods as well as effi cient separation methods are needed to analyze air samples (Barro et al. , 2009 ).

Continuously working analytical devices like the fl ame ionization detector ( FID ), photo - ionization detector ( PID ), photo - acoustic detector ( PAD ) or ion trap mass spectrometer offer a high time resolution, but often lack the required sensitivity and selectivity. Therefore, discontinuous techniques with a sample preconcentra-tion step during or after the sample collection are still preferred, especially regard-ing toxic substances where detection limits of less than 10 μ g/m 3 are demanded. Using discontinuous sampling, the sensitivity can easily be increased by a factor of 1000 to 100 000 by passing an air sample volume of 1 to 100 l through an appro-priate ‘ trap ’ where the organic ingredients are retained. In general, there are three possibilities for enriching the ingredients of an air sample:

• absorbing target compounds in a suitable liquid; • condensing target compounds at low temperatures (cryo - trapping); • adsorbing target compounds on a porous solid material.

Liquid absorption is a common technique for enriching compounds in reactive liquids like solutions of dinitrophenylhydrazine ( DNPH ) (for aldehydes), acetyl acetone (for formaldehyde) or aqueous carbonate solutions (for organic acids), both procedures which combine trapping and derivatization of the target com-pound. Another possibility is the use of dissolved alkali or acids to trap certain substances by the formation of salts in the solution.

Cryo - trapping is often used in combination with solid sorbents and therefore is less important as a stand - alone sampling method.


3

4 1 Application of Solid Sorbents for the Sampling of Volatile Organic Compounds in Indoor Air

Solid sorbents play an important role for the determination of VOCs in indoor air. They overcome some serious disadvantages of liquid absorbents:

• The direct analysis of the absorption liquid will not normally give the desired sensitivity, so additional steps are required to concentrate the target substances. These may lead to a loss of the more volatile compounds present in the sample. Moreover, the possibility of contamination, for example with phthalates, cannot be excluded when handling with glassware.

• The injection of a solvent into the gas chromatograph makes it diffi cult to identify and quantify the compounds which are eluted close to the solvent retention time. In addition to that, impurities of the solvent may interfere with the sample compounds and therefore make the quantifi cation more error - prone.

• The use of derivatization agents may increase the sensitivity, but also makes the method specifi c to a small range of compounds.

• Preconditioning of a solid phase is much easier to achieve than the purifi cation of a liquid phase. Therefore problems with blanks are reduced.

• The handling of solid sorbents, often used in pre - packed tubes, is more convenient than the use of organic solvents to trap VOCs from the air, especially for sampling in the fi eld. The typical laboratory safety procedures required for the handling of organic solvents are not required for solid sorbent samples.

If solid sorbents with a low affi nity to water are chosen, they even overcome the main disadvantage of cryo - trapping: Water gets trapped in the device and often leads to serious analytical problems as well as mechanical problems if ice is formed in the trap.

Of the variety of solid sorbents presently available Tenax, 1) activated charcoal and Carbotrap 1) are the most widely used ones. This is mostly due to the versatility they offer especially for sampling of VOCs typically found in indoor air (C6 to C16).

1.2 Solid Sorbents – A Brief Overview

Three general types of solid sorbents are mainly used for trapping VOCs in air: inorganic sorbents like silica gels or molecular sieves, carbon - based porous materials and porous organic polymers.

The main types of inorganic sorbents are silica gels, molecular sieves/zeolites, aluminum oxides and magnesium silicates. Carbon - based sorbents include acti-vated charcoals, carbon blacks, graphitized carbon blacks and graphitized molecu-lar sieves. Styrene – divinylbenzene copolymers, ethylvinylbenzene/divinylbenzene

1) For trademarks of commercial products see remarks in Table 1.1 .

1.2 Solid Sorbents – A Brief Overview 5

copolymers, polyvinylpyrrolidone, polyphenylene oxides and polyurethane foams represent the most widely used polymer sorbents. Table 1.1 shows some properties of common sorbents.

The surface area of a sorbent infl uences the amount of a given substance that can be adsorbed by the medium, whereas the surface polarity determines the general type of compounds a sorbent can be used for. The sorbents offer different suitability for VOC analysis depending on the type and amount of substance to be sampled. The inorganic sorbents are often used to trap hydrocarbons and poly-chlorinated biphenyl s ( PCB s). Carbon - based sorbents with large surface area and molecular sieves are useful to trap very low - boiling compounds, but are unsuitable for higher boiling substances; labile or reactive compounds may even decompose on active sites on the sorbent surface. Porous polymers with a comparatively small surface area allow the adsorption and desorption of high - boiling compounds like glycols, phthalates and also the trapping of reactive substances like aldehydes or acrylates. On the other hand it is diffi cult to sample low - boiling compounds like C2 - C5 - alkanes. A special case of the porous polymer sorbents is polyurethane foam ( PUF ), which can be used to collect large air samples up to 100 m 3 (Ligocki and Pankow, 1985 ). PUF is often used to trap high - boiling organopesticides like lindane or permethrine.

The different characteristics of the presented sorbents show the need to carefully choose the right adsorption medium for a given VOC mixture.

Several workers have examined the suitability of the available sorbents for general VOC analysis. Figge et al. (1987) tested the retention volumes of 26 differ-ent adsorbents using a VOC mixture of 29 compounds with boiling points from 21 ° C to 361 ° C. They rated the sorbents in four groups with decreasing sorption strength. Some of the carbon - based sorbents like Carbosieve SII showed the overall highest ability to retain organic compounds. Adsorbents with good reten-tion properties for higher boiling compounds regardless of their polarity were, Porapak Q, Chromosorb 106 and XAD - 4. The third group included Porapak S, Tenax GC and Carbopack B, which showed good retention properties only for higher boiling, non - polar VOCs. For PTFE, Chromosorb T and other weak sor-bents only poor retention volumes were found.

Brown (1996) reported the results of a multi - laboratory study concerning the suitability of different sorbents for the measurement of VOCs in the workplace environment. In this study 20 test compounds were used and the sorbents had to fulfi ll a number of test criteria. Chromosorb 106 was found to be the most versatile sorbent and especially useful for the sampling of very volatile and polar com-pounds. Other sorbents which satisfi ed the acceptance criteria were Carbotrap, Tenax TA, Tenax GR and Carbopack B.

Rothweiler, W ä ger and Schlatter (1990) and De Bortoli et al. (1992) compared two widely used adsorbents: Tenax TA and Carbotrap, regarding their performance for VOC sampling. Both groups found the two sorbents to be convenient for the sampling of non - polar organic compounds regarding background emission and sample recovery. They both agree in reporting signifi cant analyte losses when using Carbotrap with more reactive compounds like terpenes, aldehydes or acrylates.

6

1 Application of Solid Sorbents for the Sam

pling of Volatile O

rganic Com

pounds in Indoor Air

Table 1.1 Properties of some solid sorbents according to Quintana, Uribe and Arbeloa (1992) , ECA (1994) , Zielinska and Fujita (1994) , Figge, Rabel and Wieck (1987) and ISO (2001) .

Type Structure Surface area (m 2 /g)

Products Desorption Compounds tested (Starting at b.p.)

Polarity Thermal stability

Water affi nity

Inorganic Silica gels 1 – 30 Volasphere, Florisil Solvent PCBs, pesticides High ∼ 400 ° C High

Molecular sieves 500 – 800 Solvent Permanent gases High < 400 ° C High

Aluminum oxides ∼ 300 Alumina F1 Solvent Hydrocarbons High 300 ° C High

Carbon based

Activated Charcoal 800 – 1200 Solvent Non - polar and slightly polar VOCs ( > 50 ° C)

Medium > 400 ° C High

Carbon molecular Sieves

400 – 1200 Carbosieve, Ambersorb, Spherocarb Carboxen

Solvent/Thermal

Non - polar and slightly polar VOCs ( > − 80 ° C)

Low > 400 ° C Low – medium

Graphitized carbon blacks

12 – 100 Carbotrap, Carbopack, Carbograph

Thermal Non - polar VOCs ( > 60 ° C)

Low > 400 ° C Low

Porous polymers

Styrene, divinylbenzene or polyvinylpyrrolidone polymers

300 – 800 Porapak Q/N, Chromosorb 106/102,

Thermal/solvent

Non - polar and moderately polar VOCs ( > 40 ° C)

Variable < 250 ° C Low

Phenylphenylen oxide polymers

20 – 35 Tenax Thermal Non - polar VOCs ( > 60 ° C)

Low < 350 ° C Low

PU - Foams Solvent Pesticides Low < 200 ° C Low

Remarks: Tenax ® is a registered trademark of Buchem B.V., NV, NL; Carbotrap ® , Carbopack ® , Carbograph ® , Carbosieve ® and Carboxen ® are registered trademarks of Sigma - Aldrich Co., USA; Chromosorb ® is a registered trademark of Johns - Manville Corp, USA; Porapak ® is a registered trademark of Waters Associates Inc., USA; Spherocarb ® is a registered trademark of Analabs Inc., USA.; Volasphere ® : E.Merck KGaA, Germany; Florisil ® is a registred trademark of U.S. Silica Co., USA.

1.3 Active or Passive Sampling 7

Measurements of the thermal desorption effi ciencies of different sorbents were done by Cao and Hewitt (1993) . They also noticed the loss of terpenes on Carbotrap.

In the very comprehensive review article of Matisov á and Skrab á kov á (1995) the suitability of various carbon - based sorbents is discussed.

1.3 Active or Passive Sampling

Two different sampling strategies for collecting air samples exist: actively drawing air through a sorbent - fi lled cartridge or tube, or passively letting the compounds penetrate a well - defi ned sorbent bed simply through gradient - driven diffusion. The active sampling process is obviously a faster way to take an air sample and is, especially in cases where a high sensitivity and therefore a large sample volume is needed (e.g., sampling of pesticides), the recommended method. The accuracy of the sample is clearly determined by the sample volume, and measuring the correct air volume is one of the most important quality issues for the whole analy-sis when doing active sampling. Calibrated pumps, mass fl ow controllers or pumps in combination with appropriate air meters are often used to control the sample volume. If the sample volume collected is small and sampling time short, active sampling allows the measuring of dynamic processes with a higher sensitiv-ity than continuously working analytical devices (FIDs, PADs). Therefore, it is until now still the preferred method for the study of kinetic experiments in emis-sion test chambers (although with instruments like direct - inlet MS and proton transfer reaction - mass spectrometry ( PTR - MS ) becoming available this is likely to change in the future).

The use of a passive sampler is characterized by long sampling times, which are needed to allow the airborne compounds to enter the sampler and become trapped on the sorbent surface. Sampling times often exceed several days. The long sampling times show a pitfall of this sampling technique: because the sampler is in contact with the air to be sampled for a long time, very volatile compounds may have the chance not only to enter the sampler, but also to leave it on the same way, if their interaction with the sorbent is low. In this case an underestimation of the very volatile substances can be expected.

Passive samplers are not useful to monitor peak concentrations, but offer a convenient way to regard for example, long term exposure of persons in an indoor environment. A more detailed description of the passive sampling technique is given in Chapter 3 .

The decision whether active or passive sampling should be used strongly depends on the type of experiment to be carried out: to measure the mean con-centration over a long period of time in a given environment, passive sampling often is the easier way. Lewis et al. (1985) even managed to collect an analytically suffi cient sample with a special Tenax - based passive sampler within only one hour. For the determination of indoor air concentrations during fast, dynamic processes active sampling normally is preferred.


1.4 Thermal Desorption or Solvent Extraction

After collection on a sorbent the trapped VOCs need to be transferred into the analytic device, which often is a gas chromatography ( GC ) or a high performance liquid chromatography ( HPLC ) system.

Depending on the activity of the sorbent and the characteristics of the analyte there are two general possibilities of removing the sample from the sorbent. One of them is solvent extraction, which is commonly used for the inorganic sorbents as well as for PUF and the carbon - based sorbents with high surface activity. For the extraction a suitable organic solvent is selected. Dimethylformamide, carbon disulfi de (preferred because it has a low response on a FID), dichloromethane or other polar solvents, often combined with a desorbing agent like methanol or water, are used to recover the trapped substances from the adsorbent. The solvent as well as the desorbing agent need to be carefully chosen; ideally they should not interfere with any of the sample peaks (therefore their purity should be high and their peak on the used separation column should be narrow); they should not react with the sample or the sorbent; their affi nity to the sorbent should be high enough to remove the sample compound completely and their response on the used detec-tor should be low.

Thermal desorption, on the other hand, makes use of the fact that the ability of a sorbent to retain compounds dramatically decreases at elevated temperatures. Therefore, heating of the sorbent under a continuous stream of an inert carrier gas can be used to transfer the adsorbed compounds into the GC system. In general the desorption temperature should be at least 20 ° C above the boiling point of the adsorbed compounds, so the range of compounds which can be analyzed with thermal desorption is limited by the thermal stability of the sorbent. In order to avoid a loss of analyte due to decomposition on the sorbent ’ s surface at elevated temperatures the surface activity needs to be carefully chosen.

The main advantage of thermal desorption is the greater sensitivity and the absence of a solvent peak, which could interfere with a detector (i.e., mass spec-trometer ( MS )) or mask potential analyte peaks. Modern thermal desorption units allow the use of multi - bed tubes and traps (see following paragraph).

A method for trapping VOCs on Tenax adsorption tubes with subsequent thermal desorption has been standardized internationally (ISO, 2004 , 16000 - 6).

1.5 Sampler Design

A simple sampler usually consists of a glass or stainless steel tube fi lled with a certain amount of the sorbent. Typically, 50 – 500 mg of the sorbent are used. To retain the sorbent in the tubes either glass/quartz wool or small sieves can be used. Both techniques introduce a disadvantage into the sampling system: quartz wool is normally too brittle to be fi tted successfully into the tube, and small parts of the

1.5 Sampler Design 9

fi bers may be released from the tube. Silanized glass wool does not feature that disadvantage, unfortunately at thermal desorption temperatures above 250 ° C the silanization agent can be released and an increased blank value of siloxanes and similar substance can often be found in the chromatograms. In worse cases, silicon oligomers can coat parts of the thermal desorber tubing or transfer lines and may lead to decreased performance of the instrument.

Stainless steel sieves, which can be fi tted into a range of stainless steel sorbent tubes, are usually easier to handle than glass/quartz wool. It is, however, their disadvantage that some very labile compounds may degrade in contact with the metal under thermal desorption conditions. In addition, the sieves will often not completely retain the fi nes fraction of the used sorbents: this is particularly prob-lematic for the carbon - based sorbents, which are more brittle than the polymers and can therefore be crushed to fi ne particles by the thermal stress during use of a tube. The presence of a dark residue on the fi lters inside the thermal desorption unit is an indication of carbon - based sorbent migration from the tubes.

When a sorbent is to be selected for a certain sampling purpose, a number of details need to be addressed:

• Is the sorbent able to retain the targeted analyte at the given conditions (temperature, sample fl ow rate, target compound concentration etc)?

• What other compounds will be retained during that process, and can they have an adverse effect on the analysis?

• Will the sorbent material be able to withstand the sampling and analysis conditions?

To overcome problems with sorbents either being too weak to retain the very vola-tiles or too strong to allow for a desorption of the least volatile compounds of a mixture, multi - bed tubes can be used. In such a tube different sorbents of increas-ing surface area and activity are combined (Figure 1.1 ). They can be used to trap substances within a wide boiling range. High - boiling compounds get trapped in the fi rst zone of the adsorbent bed (often a porous polymer), the more volatile substances pass through that zone and are collected on an sorbent with greater activity. Therefore, the sampling direction must be accurate under all circum-stances. Desorption, in contrast, must occur in the reversed direction (back - fl ush)

Figure 1.1 A multibed sorbent tube.


Figure 1.2 Schematic of a thermal desorption process with tube desorption and normal/backfl ush trap desorption steps.

to prevent the high boiling compounds from getting in contact with the more active sorbent.

Multi - bed tubes can be a perfect solution to analyze a wide range of compounds in known samples. For air samples of unknown composition it is important to consider the possible interactions of two or more sorbents with unexpected com-pounds in that sample. Therefore, this type of tube is not in widespread use for explorative analysis.

If the thermal desorption unit is able to provide inverted (back - fl ush) gas fl ow during trap desorption (Figure 1.2 ), the same technique described for tubes can also be used for the cold trap. A multi - bed trap can considerably extend the analytical window of the instruments. Commercial cold - traps packed with quartz beads, quartz wool and Tenax TA are reported to cover a substance range from C6 to C40.

1.7 Safe Sampling Volume 11

1.6 Breakthrough Volumes

The ‘ breakthrough volume ’ is defi ned as the volume of gaseous sample that can be drawn through a sample tube before an analyte is eluted from the tube. Every sorbent has a limited capacity for a given analyte which depends on the characteristics of the sorbent, on the type of compound to be trapped and on certain sampling parameters, like the temperature and the humidity of the air (Bertoni et al. , 1981 ; Brown and Purnell, 1979 ). Breakthrough of a substance can occur if:

• the sampling speed is too high and the compounds are fl ushed through the tube without enough time to interact suffi ciently with the sorbent surface;

• the concentration is far too high, the sorbent surface gets saturated with the com-pound and the excess compound passes the sorbent without adequate retention;

• the retention ability for the given amount of substance is not suffi cient, the compound is retained, but eluted again in the ongoing sampling procedure.

While the fi rst two cases can easily be avoided by either using a low sampling rate (100 ml/min for Tenax tubes and 1 – 2 l/min for activated charcoal tubes are usually considered a safe sampling speed) or by using small enough sample volumes (Tenax: 1 – 6 l, activated charcoal: 5 – 100 l), the third case can only be overcome by selecting another sorbent or carefully choosing the right sampling temperature and a small sampling volume.

Breakthrough volumes can be determined by a direct method (monitoring the effl uent during sampling air with known concentration of a test compound), or by an indirect method, where the sorbent tube is used as a GC column and the retention time of an injected compound is used to calculate the breakthrough volume (Brown and Purnell, 1979 ).

Figure 1.3 shows the result of an experiment illustrating the breakthrough of volatile terpenes on Tenax TA: the distribution of analytes in a special adsorbent tube fi lled with Tenax TA in fi ve zones, each separated with a silanized glass wool plug, is determined. After spiking with high concentrations of different terpenes in methanol and sampling two liters of air the content of each Tenax zone was analyzed separately. The effect of chromatography - in - the - tube is clearly visible: The highest boiling compound (longifolene) is completely retained in the fi rst zone, the low - boiling compounds ( α - pinene and β - pinene) have already partially left the tube. While the β - pinene ‘ peak ’ is still located inside the tube (zone 4), the α - pinene concentration rises toward the rear end of the tube which indicates that major amounts of the substance have already left the tube.

1.7 Safe Sampling Volume

The ‘ safe sampling volume ’ ( SSV ) is usually defi ned as 70% of the 5% break-through volume (ISO, 2001 , 16017 - 1). If, for a given air sample, the sample volume


is lower than the SSV of the lowest - boiling compounds in that sample it can be assumed that all compounds were quantitatively retained in the tube. The deter-mination of SSVs is a time - consuming procedure (Martin et al. , 2003 ). Fortunately, a number of SSVs have been published for different combination of sorbents and target compounds (e.g., Ventura, Pr í hoda and Chur á cek, 1995 ). A very helpful resource in this respect is the list of breakthrough volumes of numerous com-pounds depending on the sampling temperature published on the Scientifi c Instrument Services website (Sisweb, 1996 ).

For specifi c purposes, parallel testing with two different techniques helps to reveal systematic over/underestimation of certain compounds (Salthammer and Mentese, 2008 ).

1.8 Artifacts and Interferences

1.8.1 Water Affi nity – A Chromatographic Problem

The sampled air inevitably contains a certain amount of water. If water is retained by the sorbent used for sampling it can cause severe problems during gas chro-matographical analysis: Formation of ice can lead to a clogging in ‘ purge & trap ’

Figure 1.3 Distribution of volatile compounds on an adsorbent tube after spiking and sampling 2 liters of air with high terpene concentration. The sampling direction is marked with an arrow.

1.8 Artifacts and Interferences 13

samplers and other cryo - focusing units. Water gives a high background noise level in GC – MS chromatograms and can thoroughly infl uence the signal of other detec-tors. It is also known to damage fused silica columns due to a hydrolysis of silicon fi lm material at elevated temperatures.

Therefore, the solid sorbent should ideally have a low affi nity to water, a char-acteristic easily met by the porous organic polymers. In contrast, the carbon - based sorbents and molecular sieves as well as some inorganic sorbents show a com-paratively high uptake of water. If such sorbents are used in high - humidity envi-ronments special measures have to be employed to remove water during or after sampling. Possible methods are:

• dry - purging of the sampler with helium at low temperatures (loss of more volatile organics possible);

• increasing the sampling temperature (Gawrys et al. , 2001 ); • pre - drying of the sample gas using for example, a Nafi on - dryer.

Figure 1.4 shows a scheme to ease the selection of a proper sorbent and desorption considering the target compound range and the humidity of the air to be sampled.

1.8.2 Sorbent Degradation Products and Sorbent Background

Nowadays solid sorbents can be bought pre - cleaned and less effort is required before the fi rst use. Nevertheless, all sorbents have to be cleaned thermally or by extraction to remove contaminants that formed during production, shipping or storage. For inorganic and carbon - based sorbents this is most easily accomplished by heating

Figure 1.4 Simple selection scheme for a suitable sorbent under different conditions: Boiling point range, humidity of the air sample, desorption type. Please see text for other considerations (artifact formation, surface activity).


in an inert gas stream (He, N 2 ) for some hours, an additional deactivation step may be required for silica gels, magnesium silicates or aluminum oxides depending on the favored use. Tenax can be thermally cleaned as well (280 – 300 ° C, 1 h in He or N 2 stream), but the purity of the inert purge gas should be high, otherwise artifact formation can happen with this sorbent if traces of oxygen are present.

Most of the other polymers show only limited thermal stability, so it is recom-mended that they be cleaned by solvent extraction and dry purging with pure nitrogen.

After cleaning, the samplers should be stored tightly sealed and protected from light. Refrigerated storage of the unused samplers is in most cases not required and should be avoided since standard refrigerators can be sources of contamina-tions. If cold storage is required between sampling and analysis, special attention should be directed to proper sealing of the tubes.

Almost all of the presented solid sorbents show a kind of background emission after a certain period of time. This is due to thermal or photochemical degradation of the sorbent material. Moreover, many sorbents can react with certain adsorbed compounds or reactive gases like O 2 , O 3 or NO x and reaction products will show up in the analysis (Hanson et al. , 1981 ; Clausen and Wolkoff, 1997 ; Klen ø et al. , 2002 ).

As long as the degradation products of a used sorbent are known and do not interfere with the collected sample compounds, the sorbent degradation itself is not an analytical problem. But often, especially if the compounds to be sampled are not known in advance, it is diffi cult to determine whether, for example, benzene is a blank of the used sorbent Tenax TA or a compound trapped from the air sample. Figure 1.5 shows a chromatogram obtained from a glass sampling tube fi lled with Tenax TA and stored in daylight for two weeks. It presents the important substances to be expected as Tenax blanks. Helmig (1996) suggests methods to avoid such sorbent background peaks. Other adsorbents show a spe-cifi c background as well, Tirkkonen, Mroueh and Orko (1995) present a compila-tion of compounds that were found under thermal desorption conditions for seven carbon - based and porous polymer sorbents.

Figure 1.6 presents an example for the formation of a reaction product on the surface of a sorbent. The result of a solvent extraction of activated charcoal with CS 2 and CS 2 /methanol, respectively, shows quite different artifact formation after a 48 - hour storage period. Four new compounds could be found in the sample contain-ing methanol as a desorption agent. All of them could be identifi ed as substances formed by the reaction of CS 2 and methanol in the presence of activated charcoal.

The blank problem becomes more severe when volatile organic compounds are sampled in the presence of reactive gases. Pellizzari et al. (1984) and Zielinska et al. (1986) showed the infl uence of different reactive inorganic gases on the decomposition of Tenax GC and showed possibilities to protect the adsorbent with mild reduction agents. Clausen and Wolkoff (1997) tried to use the amount of degradation products formed on Tenax as an indicator for the presence of reactive species in indoor air. Helmig (1997) reviews techniques to reduce the adverse effect of ozone on sampled VOCs.

1.8 Artifacts and Interferences 15

Figure 1.5 Background of Tenax TA after cleaning and a 2 - week storage period. Ethoxyethyl acetate is emitted from the tube cap O - ring.

While for most routine applications (e.g., chamber tests, fi eld sampling in build-ings) solid sorbent tubes can be used without having to consider most of these points, it is still important to emphasize the need to tailor the sampling tubes for more special cases, for example, during chamber testing of electronic devices or air cleaners, or during fi eld sampling at sites exposed to sunlight and/or traffi c. In the fi rst case additional sampling with carbon - based tubes might be required to avoid ozone - related artifacts (benzene, benzaldehyde) from Tenax TA. In the latter case, the presence of both reactive gases (ozone, NO x ) and labile compounds (terpenes, unsaturated hydrocarbons, aldehydes) requires some pilot testing to fi nd out which sorbent delivers the best performance under the given conditions.

1.8.3 Target Compound Degradation and Artifact Formation

Whenever organic compounds of limited stability are exposed to high temperature, reactive chemicals (liquids or gases) or active surfaces, degradation and the formation of artifacts is possible. When employing solid sorbents the sampling/storage step as well as the extraction/analysis step may lead to degradation of analytes. Coeur et al. (1997) reported the degradation of α - pinene and sabinene on Tenax and Carboxene. Several other terpenes (camphene, limonene, cymene, terpinolene) were formed as artifacts. The degradation of organo - sulfurous com-pounds during thermal desorption was studied by Baltussen et al. (1999) for several sorbents.


Figure 1.6 Background of activated charcoal after extraction (a) with CS 2 and (b) CS 2 /Methanol. 1, 2 and 3 are condensation products of CS 2 and methanol.

Especially when long - term storage of samples is required the monitoring of analyte degradation is of special interest (Volden et al. , 2005 ).

1.9 Conclusions

Solid sorbents offer a convenient way to enrich the contents of an air sample. The variety of sorbents currently available allows the sampling of gaseous compounds from VVOCs to SVOCs. However, none of the existing sorbents is capable of

References 17

retaining all compounds, so either a combination of adsorbents with different characteristics (multi - bed tubes) or one sorbent specially chosen for the actual analytical problem has to be used.

Although new techniques of analyzing VOCs like sensor systems and online mass spectrometers have been introduced in the last decade, the use of solid sor-bents in combination with thermal desorption and gas chromatographic separa-tion still seems to be the preferred method due to the sensitivity, selectivity, convenience of use and the reliability that has been proven over the years.

References

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Barro , R. , Regueiro , J. , Llompart , M. and Garcia - Jares , C. ( 2009 ) Analysis of industrial contaminants in indoor air: Part 1. Volatile organic compounds, carbonyl compounds, polycyclic aromatic hydrocarbons and polychlorinated biphenyls . Journal of Chromatography A , 1216 , 540 – 66 .

Bertoni , G. , Bruner , F. , Liberti , A. and Perrino , C. ( 1981 ) Some critical parameters in collection, recovery and gas chromatographic analysis of organic pollutants in ambient air using light adsorbents . Journal of Chromatography , 203 , 263 – 70 .

Brown , R.H. ( 1996 ) What is the best sorbent for pumped sampling - thermal desorption of volatile organic compounds? Experiences with the EC sorbents project . Analyst , 212 , 1171 – 5 .

Brown , R.H. and Purnell , C.J. ( 1979 ) Collection and analysis of trace organic vapour pollutants in ambient atmospheres . Journal of Chromatography , 178 , 79 – 80 .

Cao , X. - L. and Hewitt , C.N. ( 1993 ) Thermal desorption effi ciencies for different adsorbate/adsorbent systems typically used in air monitoring programs . Chemosphere , 27 ( 5 ), 695 – 705 .

Clausen , P.A. and Wolkoff , P. ( 1997 ) Degradation products of Tenax TA formed during sampling and thermal desorption

analysis: indicators of reactive species indoors . Atmospheric Environment , 31 ( 5 ), 715 – 25 .

Coeur , C. , Jacob , V. , Denis , I. and Foster , P. ( 1997 ) Decomposition of a - Pinen and sabinen on solid sorbents, Tenax TA and Carboxen . Journal of Chromatography A , 786 , 185 – 7 .

De Bortoli , M. , Kn ö ppel , H. , Pecchio , E. , Schauenburg , H. and Vissers , H. ( 1992 ) Comparison of Tenax and Carbotrap for VOC sampling in indoor air . Indoor Air , 2 , 216 – 24 .

ECA ( 1994 ) Sampling strategies for volatile organic compounds (VOCs) in indoor air . report 14, EUR 16051 EN Offi ce for Offi cial Publications of the European Community, Luxembourg.

Figge , K. , Rabel , W. and Wieck , A. ( 1987 ) Adsorptionsmittel zur Anreicherung von organischen Luftinhaltsstoffen . Fresenius ’ Zeitschrift f ü r Analytische Chemie , 327 , 261 – 78 .

Gawrys , M. , Fastyn , P. , Gawlowski , J. , Gierczak , T. and Niedzielski , J. ( 2001 ) Prevention of water vapour adsorption by carbon molecular sieves in sampling humid gases . Journal of Chromatography A , 933 , 107 – 16 .

Hanson , R.L. , Clark , C.R. , Carpenter , R.L. and Hobbs , C.H. ( 1981 ) Evaluation of Tenax - GC and XAD - 2 as polymer adsorbents for sampling fossil fuel combustion products containing nitrogen oxides . Environmental Science and Technology , 15 ( 6 ), 701 – 5 .

Helmig , D. ( 1996 ) Artefact - free preparation, storage and analysis of solid adsorbent sampling cartridges used in the analysis of volatile organic compounds in air . Journal of Chromatography A , 732 , 414 – 17 .


Helmig , D. ( 1997 ) Ozone removal techniques in the sampling of atmospheric volatile oranic trace gases . Atmospheric Environment , 31 ( 21 ), 3636 – 51 .

ISO ( 2001 ) 16017 - 1 . Indoor, Ambient and Workplace Air – Sampling and Analysis of Volatile Organic Compounds by Sorbent Tube/Thermal Desorption/Capillary Gas Chromatography – Part 1: Pumped Sampling , International Organization for Standardization , Geneva, Switzerland , pp. 1 – 28 .

ISO ( 2004 ) 16000 - 6 . Indoor Air – Part 6: Determination of Volatile Organic Compounds in Indoor and Test Chamber Air by Active Sampling on Tenax TA ® Sorbent, Thermal Desorption and Gas Chromatography Using MS/FID , International Organization for Standardization , Geneva, Switzerland .

Klen ø , J.G. , Wolkoff , P. , Clausen , P.A. , Wilkins , C.K. and Pedersen , T. ( 2002 ) Degradation of the adsorbent Tenax TA by nitrogen oxides, ozone, hydrogen peroxide, OH radical and limonene oxidation products . Environmental Science and Technology , 36 , 4121 – 6 .

Lewis , R.G. , Mulik , J.D. , Coutant , R.W. , Wooten , G.W. and McMillin , C.R. ( 1985 ) Thermally desorbable passive sampling device for volatile organic chemicals in ambient air . Analytical Chemistry , 57 , 214 – 19 .

Ligocki , M.P. and Pankow , J.F. ( 1985 ) Assessment of adsorption/solvent extraction with polyurethane foam and adsorption/thermal desorption with Tenax GC for the collection and analysis of ambient organic vapors . Analytical Chemistry , 57 , 1138 – 44 .

Martin , N.A. , Marlow , D.J. , Henderson , M.H. , Goody , B.A. and Quincey , P.G. ( 2003 ) Studies using the sorbent Carbopack X for measuring environmental benzene with Perkin - Elmer - type pumped and diffusive samples . Atmospheric Environment , 37 , 871 – 9 .

Matisov á , E. and Skrab á kov á , S. ( 1995 ) Carbon sorbents and their utilisation for the preconcentration of organic pollutants

in environmental samples . Journal of Chromatography A , 707 , 145 – 79 .

Pellizzari , E. , Demian , B. and Krost , K. ( 1984 ) Sampling of organic compounds in the presence of reactive inorganic gases with Tenax GC . Analytical Chemistry , 56 , 793 – 8 .

Quintana , M. , Uribe , B. and Lopez Arbeloa , J. ( 1992 ) Sorbents for active sampling , in Clean Air at Work – New Trends in Assessment and Measurement for the 1990s (eds R. Brown , M. Curtis , K. Saunders and S. Vanderschiesche ), The Royal Society of Chemistry , Cambridge, UK , pp. 124 – 34 .

Rothweiler , H. , W ä ger , P.A. and Schlatter , C. ( 1990 ) Comparison of Tenax TA and Carbotrap for sampling and analysis of volatile organic compounds in air . Atmospheric Environment , 25B ( 2 ), 231 – 5 .

Salthammer , T. and Mentese , S. ( 2008 ) Comparison of analytical techniques for the determination of aldehydes in test chambers . Chemosphere , 73 , 1351 – 6 .

Sisweb ( 1996 ) Scientifi c Instrument Services Inc., Ringoes, NJ 08551, USA, http://www.sisweb.com/index/referenc/breakthrough.htm (accessed 03 April 2009).

Tirkkonen , T. , Mroueh , U. - M. and Orko , I. ( 1995 ) Tenax as a collection medium for volatile organic compounds . NKB Committee and Work Reports 1995:06 E .

Ventura , K. , Pr í hoda , P. and Chur á cek , J. ( 1995 ) Aplication of solid sorbents to the trace analysis of alkyl esters of acrylic acid in air . Journal of Chromatography A , 710 , 167 – 73 .

Volden , J. , Thomassen , Y. , Greibrokk , T. , Thorud , S. and Molander , P. ( 2005 ) Stability of workroom air volatile organic compounds on solid adsorbents for thermal desorption gas chromatography . Analytica Chimica Acta , 530 , 263 – 71 .

Zielinska , B. and Fujita , E. ( 1994 ) Organic gas sampling , in Environmental Sampling For Trace Analysis (ed. B. Markert ), VCH Verlagsgesellschaft , Weinheim, Germany .

Zielinska , B. , Arey , J. , Ramdahl , T. , Atkinson , R. and Winer , A.M. ( 1986 ) Potential for artifact formation during Tenax sampling of polycyclic aromatic hydrocarbon . Journal of Chromatography , 363 , 382 – 6 .

Sampling and Analysis of SVOC s and POM s in Indoor Air Per Axel Clausen and Vivi Kofoed - S ø rensen

2

2.1 Introduction

The objective of this chapter is to focus on important aspects of sampling and analysis of semi - volatile organic compound s ( SVOC s) and particulate organic material s ( POM s) in indoor air and possible pitfalls. It can therefore only provide a limited overview on recent studies. The description of techniques and procedures are intended to be from a practical and not a theoretical point of view. The idea is to give information to increase the reader ’ s knowledge on physical/chemical prop-erties relevant to the measurements, and give inspiration for improving perfor-mance and reliability of measurements. The number of combinations of sampling and analytical methods is almost infi nite. Here, we limit the discussion to combina-tions that are relevant to the fi eld or have been used successfully in our laboratory. Recently, more general state - of - the - art reviews also including sampling and analy-sis of SVOCs/POMs have been published covering environmental analysis (Koester and Moulik, 2005 ), passive sampling and extraction techniques (Namiesnik et al ., 2005 ; Seethapathy, G ó recki and Li 2008 ), solid - phase microextraction (Ouyang and Pawliszyn, 2006a, 2006b ) and membrane extraction (Hylton and Mitra, 2007 ).

All the steps from preparation of sampling equipment, sampling, transport, storage, preparation of samples for analysis and the analysis itself are subject to artifact problems. Artifact problems can mean too little or too much of the analyte of interest. It can also mean the presence of a compound that in fact does not exist in the measured object, or the complete absence of a compound that in fact exists in the measured object. Artifact problems for the different steps will be discussed in the individual paragraphs.

2.2 Defi nitions and Properties of SVOC s and POM s

The abbreviations SVOC ( semi - volatile organic compound ) and POM ( particulate organic material ) have been proposed by the World Health Organization (WHO,


19

20 2 Sampling and Analysis of SVOCs and POMs in Indoor Air

1989 ). SVOCs were defi ned as compounds with boiling points ranging from 240 – 260 ° C to 380 – 400 ° C with polar compounds in the high ends of the intervals. POMs were defi ned as compounds with boiling points higher than 380 ° C. SVOCs have also been defi ned as compounds with saturation vapor pressures ranging from 10 − 2 kPa (10 − 1 torr) to 10 − 8 kPa (10 − 7 torr) (Lewis, 1986 ). Both defi nitions mean that SVOCs may be distributed between the gas phase and the particle phase in indoor air. Probably the best defi nitions are that SVOCs are found in signifi cant amounts in both the gas and particle phase and that POMs are found in signifi cant amounts only in the particle/solid phase in the indoor environment.

2.2.1 Gas/Particle Partitioning in Indoor Air

SVOCs may be simultaneously present in meaningful amounts in the gas phase and on the surface of airborne particles. Therefore it is important to sample both the gas phase and the particle phase. This distribution between the gas phase and particles is known as ‘ partitioning ’ and may roughly be illustrated by Junge ’ s equa-tion (Junge, 1977 ) derived on the basis of adsorption theory:

φ θ θ= ⋅ + ⋅( )( )c p c0 (2.1)

φ is the relative concentration of a compound in the particle phase/total air con-centration (dimensionless); c is a constant that depends on the molecular weight and heat of condensation and does not vary much with the substances involved as long as only physical adsorption is considered θ is the aerosol surface area per volume (cm 2 /cm 3 ); and p 0 is the saturation vapor pressure (mm Hg). This simple equation is based on adsorption theory only and illustrates the partitioning very well (see Figure 2.1 ). However, it can only be considered a rough approximation since organic compounds are associated with particles not only through adsorption but also through some combination of surface interactions involving van der Waals and Lewis acid/base forces, and absorption such as insertion into an organic or aqueous layer on a particle ’ s surface.

In today ’ s literature (Weschler, 2003 ) the thermodynamic particle - gas partition coeffi cient, K p (m 3 / μ g), for a given organic compound is defi ned as:

K F c c cgp p gTSP= ( ) = (2.2)

F is the equilibrium particle phase concentration of a compound (ng/m 3 ); TSP is the concentration of total suspended particles ( μ g/m 3 ); c g is the equilibrium gas phase concentration and c p is the concentration within the particle phase of the compound (ng/m 3 ). In essence K p is the ratio of the fractional concentration of a given organic compound on particles to its concentration in the gas phase. Within a given class of organic compounds, and in some cases even among different classes, experimental data shows that the logarithm of K p tends to correlate in a linear fashion with the log of the saturation vapor pressure. Using Equation 2.2

2.2 Defi nitions and Properties of SVOCs and POMs 21

and a vapor pressure of 1.9 × 10 − 10 atm for the ubiquitous plasticizer di - (2 - ethylhexyl)phthalate ( DEHP ) (Afshari et al ., 2004 ), the assumptions that TSP is 20 μ g/m 3 and the DEHP gas phase concentration is 0.07 μ g/m 3 , Weschler (2003) fi nds that 83% of the total air concentration of DEHP is in the particle phase in a typical indoor environment.

2.2.2 Surface Adsorption

The gas/particle partitioning in indoor air is a special case of the general sorption and desorption properties of organic compounds at surfaces. This has been reviewed comprehensively by Weschler (2003) in relation to the indoor environ-ment and the theory further substantiated by demonstrating that the average air-borne particle concentration of phthalates in a large number of buildings may be estimated from the measurement of the phthalate dust concentrations in the build-ings (Weschler, Salthammer and Fromme, 2008 ). In general, the lower the vapor pressure the larger is the tendency to adsorb to surfaces. It is very important to keep in mind these ‘ sticky ’ properties of SVOCs/POMs. Also under the conditions of sampling, preparation of samples, and analysis, the SVOCs/POMs will adsorb onto most surfaces. They can adsorb onto glass and stainless steel from the gas phase in emission chambers (Clausen et al ., 2004 ), to glassware and polyethylene vials from solutions (Pinto, Jos é and Cordero, 1994 ), to ‘ cold spots ’ in gas chro-matographic systems, etc. It has also been shown that gas phase SVOCs may be

Figure 2.1 Expected values of φ = the relative SVOC concentration in the particle phase/total air concentration as a function of the logarithm of the saturation vapor pressure, p 0 , in the region with particle surface areas per volume between clean air and urban air (Junge, 1977 ). The values are only a rough indication of φ as a function of log( p 0 ).


adsorbed to and/or absorbed by cotton fi bers, wallpaper, wall plaster (1 cm below the surface), furniture, carpets, and dust (Gebef ü gi, 1989, 1995 ).The opposite may also happen: samples can be contaminated from surfaces, therefore it is very important that all equipment used for measurements of SVOCs and POMs is carefully cleaned.

2.2.3 Health Related Properties

The biological effects and transport properties of particulate air - pollution are closely related and depend strongly on the particle size distribution. The size of the particles determines the deposition site and deposition effi ciency in the human respiratory system. Coarse particles deposit in the upper respiratory system (nose and throat), whereas the fi ne (PM 1 and PM 2.5 ) and ultrafi ne particles ( ≤ 100 nm) are predominantly retained in the deeper bronchio - alveolar region of the lung (Heyder et al ., 1986 ). Some ultrafi ne particles also deposit in the nose and throat owing to their high diffusion rates. Consequently, since indoor particulate air pol-lution mainly occurs in the submicrometer range, it is predicted to deposit mainly in the gas - exchange region in the deeper respiratory tract with lesser amounts in the nose and throat. Therefore, in order to relate measurements to health effects SVOCs and POMs should be sampled with size selective inlets (PM 1 or PM 2.5 ). However, it must be taken into consideration that the gas/particle partitioning may change during inhalation, equivalent to the change of partitioning observed during sampling (Lewis, 1986 ).

2.3 Compounds and Matrices in the Indoor Environment

Many objects which contain and emit SVOCs and POMs can be found in the indoor environment. SVOC can be released from ordinary furniture, building materials, cleaning products, human activity etc. Examples of SVOCs and POMs studied in indoor environments are phthalates and non - ionic surfactants (Clausen et al ., 2003 ), phosphor organic compounds (Wensing, Uhde and Salthammer, 2005 ), polycyclic aromatic hydrocarbon s ( PAH s) (Sheldon et al ., 1993 ; Jensen, Kofoed - S ø rensen and Clausen, 2005a ), brominated fl ame retardants (Sj ö din et al ., 2001 ), PCBs (Kohler, Zennegg and Waeber, 2002 ), pesticides (Krause, Chutsch and Englert, 1989 ; Barro et al ., 2006 ), musk fragrances (Fromme et al ., 2004 ), mineral oil residues (Clausen and Wolkoff, 1997b ), and fatty acids and salts (Clausen, Wilkins and Wolkoff, 1998 ). Examples of matrices are gas and particles in air, surface dust, building materials, furniture, computers and other consumer products. Some SVOCs are slowly emitted from the products (e.g., phthalates) and are common pollutants in indoor air and surface dust. Increased use of wood - burning stoves for domestic heating in Denmark has additionally re - introduced wood burning as an important PAH source in indoor environments (Glasius,

2.4 Sampling, Transport and Storage of SVOC/POM Samples 23

W å hlin and Palmgren, 2004 ). PAHs also exist as constituents in emissions from typical indoor sources such as cooking and smoking. Hence, human exposure to particulate matter and PAHs in indoor environments may be a mixture of the contributions from infi ltrated outdoor air and indoor sources, respectively.

2.4 Sampling, Transport and Storage of SVOC / POM Samples

Preparation of the equipment for sampling is a very important step and will be described in the individual paragraphs. Air samples are usually collected by active sampling on a fi lter (for the particle phase) followed by an adsorbent tube (for the gas phase) using a pump. Samples of surface dust for POM analysis are often sampled on a fi lter using a vacuum cleaner as pump. If quantitative measurements are required, it must be ensured that the components do not leave (break through) the fi lter/adsorbent during sampling and that the background levels in blank samples are low. Furthermore, it is important to prepare fi eld blanks that are as identical as possible to the real samples and transported, stored, analyzed and otherwise treated in exactly the same way. This is done in order to check that no artifact formation or other contamination has happened during these steps.

2.4.1 Preparation of Sampling and Analysis Equipment

The major problem during measurement of SVOCs/POMs probably is the reduc-tion of the background level in order to have low contamination in blanks. This means that everything used for the sampling and analysis must be cleaned and the purity must be checked by analysis of blank samples.

2.4.1.1 Background Contamination and Loss of Target Compounds The level of the background contamination is important because of the relatively low contents of target SVOCs/POMs in the fi eld samples. It may be necessary to clean and treat all the glassware, pipettes and equipments in special manners for example, rinse with acetone or bake - out at high temperatures. Also the sam-pling equipment, fi lters and sorbent tubes must be as clean as possible. If com-mercial disposable sampling equipment is used without cleaning, the background must be checked and the result of the samples corrected. The background level can be determined by analysis of 5 – 10 blank samples, stored, transported and treated as the fi eld samples themselves. Correct storing of the pure sampling equipment prior to sampling and the samples after sampling is also very impor-tant. Incorrect storing and transportation before analysis may contaminate the pure equipment and the samples or result in breakdown or loss of the target components. Some components must be protected from light or must be kept cool to retain stability. Correct storing depends on the type of components present in the samples.


2.4.1.2 Cleaning of Filters Quartz fi ber fi lters can easily be cleaned during one hour by ozone plasma in a low temperature asher whereas Tefl on membrane fi lters which cannot be heated or treated with ozone plasma are diffi cult to clean even by Soxhlet extraction (Clausen and Wolkoff, 1997b ). There is also a risk that glass fi ber fi lters will melt during high temperature cleaning. The purity of the fi lters must be checked by analyzing blank fi lters using the same procedure as for the real samples.

2.4.1.3 Cleaning of Sorbents The choice of sorbent, sampling and preparation methods includes the laboratory apparatus facilities and the purpose of the investigation.

The tubes for solvent desorption are typically constructed of glass tubing, with both ends fl ame - sealed and containing two sections of suitable sorbent. The fi rst section contains the adsorbent for the sample and the second section a back - up to test for breakthrough. Packed tubes are commercially available, alternatively empty tubes can be packed in the laboratory. The commercial sorbent tubes for solvent desorption, for example XAD glass tubes, are assumed to be clean but their cleanli-ness must, regardless of the assurances of the manufacturers, be checked by analyzing a minimum of fi ve blank tubes by the same desorption and analytical procedures as used for the real samples.

If the tubes are packed in the laboratory a procedure for cleaning tubes and sorbent material must be developed. For thermal desorption ( TD ) (see Section 2.6.1 ) the tubes should be compatible with the TD instrument. They are often constructed of stainless steel tubing and contain typically about 200 – 300 mg of porous polymer for example, Tenax TA or 300 – 500 mg carbon or graphitized carbon sorbent. The sorbents are retained by stainless steel gauzes. Before use the sorbent tubes must be cleaned and analyzed to ensure that the TD blank back-ground is suffi ciently low. Tenax TA tubes for TD can be cleaned in a stream of ca. 60 ml nitrogen per min at 275 ° C for at least one hour or cleaned in the TD system used for the analysis at the same condition as used for the samples. The tubes should be checked for contaminations by TD – GC analysis before use. If the chromatograms of the blank tubes show unacceptably high background, the tubes need to be reconditioned. TD tubes should be sealed with metal screw - cap fi ttings with Tefl on seals. The tubes can be stored in refrigerator in Rilsan bags or special containers before run. It is advisable to check the background of blank tubes if the tubes are stored for an extended period before use because even with tight seals and correct storing they can become contaminated. It is also required to test the storage durability of the samples collected on the sorbent tubes under well - defi ned conditions (time, temperature, light etc.). After sampling, the tubes should be sealed with cleaned Tefl on, not rubber, seals and/or screw - cap fi ttings.

2.4.1.4 Cleaning of Glassware and Other Equipment Smaller glassware may be washed by standard procedures in the central dish-washer and afterwards cleaned in an oxygen plasma asher for 60 minutes at 100 W (Clausen et al ., 2004 ) or by high temperature (500 ° C) in an oven (Nielsen, Clausen


and Jensen, 1986 ) and stored in cleaned aluminum foil (rinsed with acetone). Glassware such as bottles and fl asks can be rinsed with acetone and stored in clean aluminum foil. Before use, disposable glassware (micro vials) must also be cleaned and stored in cleaned aluminum foil. High temperature cleaning or solvent rinsing are methods which can be used for nearly all cleaning procedures.

2.4.2 Sampling SVOC s / POM s in Air

2.4.2.1 Filter/Sorbent Sampling When SVOCs are sampled from air, both the gas phase and the particle phase are usually collected. It is safest to use both a fi lter and an adsorbent to collect the particle phase and the gas phase SVOCs, respectively (see Figure 2.2 ). SVOCs/POMs associated with particles in the indoor environment may desorb during sampling and break through the fi lter as gas phase. On the other hand SVOCs/POMs that exist in the gas phase in the indoor environment may be adsorbed to the fi lter and/or the particles captured by the fi lter. Thus the distribution between fi lter and adsorbent does not (necessarily) refl ect the distribution in the sampled air. DEHP can be collected effectively on quartz fi ber fi lters (Clausen and Wolkoff, 1997b ) but can break through Tefl on membrane fi lters (Clausen, Raaschou - Nielsen and Bisgaard, 2002 ). However, sometimes a fi lter is not needed since SVOCs/POMs associated with particles may or may not be effectively captured by sorbent beds (Kogan et al ., 1993 ). Also, when only gas phase SVOCs are expected, as in chamber testing of emission of DEHP from polyvinyl chloride ( PVC ) (Clausen et al ., 2004 ), a fi lter is not needed. Breakthrough must always be tested even if both a fi lter and an adsorbent are used for sampling. Thus, always use a back - up adsorbent tube to check for breakthrough from the sample tube or test break-through by spiking the fi lter/adsorbent with the SVOC/POM of interest and suck clean air through the sampling device, see for example, Yoshida, Matsunaga and Oda (2004) and Elfl ein et al . (2003) .

Usually, large - volume air sampling is necessary to measure SVOCs/POMs because of the low air concentrations and dilution of the samples by the following solvent extraction. One example is sampling of PAHs in indoor air using particle size selective inlets (PM 2.5 ), glass fi ber fi lter and sample and backup adsorbent tubes (Tenax TA) at 3.5 l/min for one week resulting in 35 m 3 air (Jensen, Kofoed - S ø rensen and Clausen, 2005a, 2005b ). However, SVOCs can be measured by

Figure 2.2 Filter/adsorbent sampler with backup adsorbent tube designed for thermal desorption analysis of SVOCs (Clausen and Wolkoff, 1997b ).


low volume sampling (30 – 100 l) in indoor air using fi lter/adsorbent sampling (see Figure 2.2 ) and TD analysis (see Section 2.6.1 ) (Clausen and Wolkoff, 1997b ).

Polyurethane foam ( PUF ) offers a much lower resistance to air fl ow than granu-lar sorbents such as XAD - 2 and it is inexpensive and easy to use (Lewis, 1989 ). However, PUF cannot retain SVOCs with high vapor pressure and should be backed up by for example, XAD - 2. It has been shown that PUF quantitatively retain a series of household pesticides (Elfl ein et al ., 2003 ).

An alternative to the traditional air sampling media is the use of solid phase extraction ( SPE ) membranes that have a higher permeability than the traditional sorbent cartridges. The Empore SPE membranes consist of polytetrafl uoroethyl-ene ( PTFE ) fi brils impregnated with small particles of essentially any solid sorbent. Recent studies have shown good performance of the Empore membranes for sampling of indoor SVOCs/POMs (Yoshida, Matsunaga and Oda, 2004 ; Tollback et al ., 2006 ). Another alternative may be sorbent impregnated fi lter s ( SIF s) that are ordinary glass fi ber fi lters coated or impregnated with fi nely ground XAD resin. SIFs have been shown to be promising alternatives to PUF for sampling gas phase SVOCs (Galarneau et al ., 2006 ).

2.4.2.2 Determination of the Gas/Particle Partitioning: Denuder Sampling Probably the best method for estimating the gas/particle partitioning is denuder sampling. In principle, a denuder is a tube that is coated on the internal surface with an adsorbent suitable for the SVOC of interest. The sample air is drawn through the tube at a speed that allows the gas phase SVOCs to diffuse to the internal surface and be captured by the adsorbent while the particles continue through the tube. A fi lter/adsorbent sampler is placed at the outlet end of the denuder. This collects the particles and the gas phase SVOCs that may desorb from the particles during the sampling. The fi rst diffusion denuder sampler for SVOCs was developed by Lane et al . (1988) , however, it was not possible to recover the gas phase SVOCs adsorbed in the denuder. Later this problem was solved by use of fi nely granulated XAD - 4 adsorbent in the denuder (Gundel et al ., 1995 ). The development of denuders for sampling SVOCs is still at the research level and very complex systems are used to evaluate these samplers (Fan, Brook and Mabury, 2003 ; Lewtas et al ., 2001 ; Eatough et al . (1993) .

2.4.2.3 Artifact Formation Caused by Reactive Gases in Indoor Air Indoor air contains reactive gases such as ozone and nitrogen oxides. During sampling these gases may react with the sorption media or the sorbed SVOCs/POMs. The adsorbent Tenax reacts with for example, nitrogen oxides, ozone, hydrogen peroxide, OH radical, and limonene/ozone oxidation products (Klen ø et al ., 2002 ; Clausen and Wolkoff 1997a ) and forms a variety of products. These products may be formed in large amounts during sampling of even small air volumes (30 – 100 l) in indoor environments (Clausen and Wolkoff, 1997a ). Filter reaction artifacts have been shown to lead to an underestimation of the actual PAH content of urban air particulate matter by up 50% or more (Schauer, Niessner and


P ö shl, 2003 ) and sampling gas - phase PAHs from the ambient atmosphere onto Tenax solid adsorbent in the presence of oxides of nitrogen can lead to artifact formation of nitro derivatives via reactions within the Tenax adsorbent (Zielinska et al ., 1986 ). This artifact formation problem has not yet been solved. However, the degradation of the target compounds during sampling can be measured by spiking the sampler with deuterated standards of the target compounds and esti-mating their degradation. Alternatively, removal of the reactive gases before they enter the sampler could be attempted. Manganese dioxide - coated copper grids or potassium iodide - coated denuders may be used to remove ozone (Calogirou et al ., 1996 ).

2.4.2.4 Air Sampling Pumps Air sampling of SVOCs and POMs usually involves pumps, which must have a fl ow controller, since it is very important to know the exact sampling volume. The pumps must be calibrated and checked before and after sampling, and for long - term sampling also within the sampling period. Calibration and checking of the pumps must be done with a traceable instrument as for example an elec-tronic bubble fl ow - meter. Before sampling it must be assured that the pump is able to compensate for the pressure drop over the sampling device, and will run stably throughout the desired sampling period. Furthermore, it must be assured that the exhaust from the pumps does not contaminate the sampling space. Alternatively, the pumps must be placed outside the sampling space or the exhaust vented outside. Indoor sampling often requires low noise from the pumps. Therefore, they must be acoustically insulated, even small pumps for low fl ows (200 ml/min) (Clausen and Wolkoff, 1997b ). Quiet indoor SVOC samplers have been developed for fl ows of 20 l/min (Wilson, Chuang and Kuhlman, 1991 ), 27 l/min (Loiselle, Offermann and Hodgson, 1991 ), and 224 l/min (Wilson et al ., 1989 ). The perturbations of the sampling space due to the high fl ow sampling must be estimated, for example, increased air exchange rate if the exhaust is vented outdoors.

2.4.3 SVOC s / POM s in Surface Dust

2.4.3.1 Filter Sampling with Vacuum Cleaner The most frequently used methods for sampling indoor surface dust for SVOC/POM analysis is simply to use dust from a vacuum cleaner dust bag, see for example, Krause, Chutsch and Englert (1989) , or a special vacuum cleaner mouth - piece containing a fi lter, see for example, Ø ie, Hersoug and Madsen (1997) . Both the mouth - piece and the fi lter should be cleaned prior to use. It must be assured that the SVOCs and POMs of interest are quantitatively captured by the fi lter (e.g., phthalates are quantitatively captured by quartz fi ber fi lters (Clausen and Wolkoff, 1997b )) and that the vacuum cleaner exhaust does not contaminate the samples with for example, phthalates. However, the vacuum cleaner sampling method is probably very dependent on the sampling conditions (e.g., carpet or hard fl oor


covering and fl ow rate) and it may be diffi cult to compare different studies or even different samples within the same study.

2.4.3.2 Specially Designed Dust Sampler To avoid the problems with vacuum cleaner surface dust sampling a special surface dust sampler, HVS - 3, was developed (Roberts et al ., 1991 ). The sampler consists of a nozzle that can be adjusted to a well defi ned distance to the surface, a cyclone that collects the dust particles, an air pump, and an exhaust fi lter to capture particles that are not retained in the cyclone. SVOCs may break through the cyclone as vapors and a PUF plug can be inserted after the cyclone. HVS - 3 has been modifi ed to ensure a more constant suction pressure and volume, and a known sampled area (Gyntelberg et al ., 1994 ). The design and use of the HSV3 has now been standardized (ASTM, 1997 , D 5438 - 94).

2.4.4 SVOC s / POM s in Building Materials and Consumer Products

2.4.4.1 Indoor Material Samples Containing SVOC s / POM s All kinds of building materials, furniture, textiles, computers etc. are potential sources of SVOCs/POMs and estimation of the total content in the material may be required. One reason to estimate the total or initial content is that it is an important parameter in physical based emission models for example, for phthalate emission from vinyl fl ooring (Clausen et al ., 2007 ). The materials can be purchased as new, or sampled from the indoor environment under investigation. The materi-als should be stored in a way that preserves their content of SVOCs/POMs and prevents contamination and degradation of the materials and their content of SVOCs/POMs. The content of SVOCs/POMs in the sampled materials can be estimated by extraction (see Section 2.5.1 ).

2.4.4.2 Testing Emission of SVOC s from Indoor Materials in Chambers Small testing chambers such as the FLEC ( Field and Laboratory Emission Cell ) and the CLIMPAQ ( Chamber for Laboratory Investigations of Materials, Pollution and Air Quality ) are good alternatives to large chambers to study the emission of SVOCs from building materials. Emission testing is important for identifying sources of specifi c SVOCs in air and dust samples from an indoor environment under investigation. If the air in the chamber is particle free, sampling can be done without fi lter. The number of studies of emission of SVOCs from indoor materials published in peer - reviewed journals is few (Clausen et al ., 2004, 2007 ; Afshari et al ., 2004 ; Uhde et al ., 2001 ; Clausen 1993 ). This is probably due to dif-fi culties with the analysis of the very low concentrations and the fact that long - term studies are necessary in order to get equilibrium in the chambers.

Emission testing of DEHP from a PVC fl ooring has been studied for up to 472 days in both the FLEC and the CLIMPAQ (Clausen et al ., 2004 ). The loading of the CLIMPAQs was varied but was constant in the FLECs. The sorption properties of the FLEC and the CLIMPAQ were investigated using different methods. In


addition, the uptake of DEHP by offi ce fl oor dust soiled on the PVC fl ooring was studied in a CLIMPAQ experiment. The concentration versus time curves in both the FLECs and the CLIMPAQs increased slowly over about 150 days and reached a quasi - static equilibrium at 1 μ g · m − 3 . The main conclusions were that the emis-sion rate of DEHP was limited by gas phase mass transport and the dust layer increased the emission rate by increasing the external concentration gradient above the surface of the PVC. These conclusions were based on the facts that the specifi c emission rate was inversely proportional to the loading and that the dust had sorbed about four times as much DEHP over a 68 - day period as emitted in the gas phase experiments (see Figure 2.3 ). About one half of the emitted DEHP was deposited on the internal surfaces of both the FLEC and the CLIMPAQ. The consequence of the gas phase transport limited emission is that the emission rate is very sensitive to the environmental conditions and thereby depends on the specifi c indoor environment that the material is placed in.

To be able to predict the release of SVOCs from a material to the indoor envi-ronment it is important to understand the fundamental mechanisms in order to mathematically model the emissions. The emission behavior of DEHP from PVC in the FLEC and CLIMPAQ experiments (Clausen et al ., 2004 ) have now been successfully modeled (Xu and Little, 2006 ). Fluid building materials such as paints (Clausen, 1993 ; Xu and Little, 2006 ) and wood oil (Clausen, 1997 ) may also emit SVOCs and are usually used on large indoor surfaces such as walls, ceilings and fl oors. Such wet materials may be applied on substrates like wood or plaster board. The emission of for example, Texanol from water - based paint was found likely to be limited by gas phase mass transport (Clausen, 1993 ) similar to the DEHP emis-sion from PVC (Clausen et al ., 2004 ).

Figure 2.3 Chamber gas phase and dust concentrations of DEHP in fl oor dust soiled on PVC fl ooring in CLIMPAQs for three different scenarios. The bar graphs show the dust concentrations (error bars are analytical 95% confi dence intervals). The curves show

the corresponding gas phase concentrations. Error bars of the gas phase concentrations are omitted for clarity. The average relative standard deviation was 25% (Clausen et al ., 2004 ) .


2.5 Preparation of SVOC/POM Samples for Analysis

One key issue of a study is to minimize sampling and analytical errors, loss of target compound and other artifacts through the preparation steps. Therefore it is important to check all steps in the method. Preparation of samples collected on fi lters and sorbents, dust samples, and material samples will be discussed here. In Section 2.6 some analysis techniques that require no or minimal sample prepa-ration will be mentioned briefl y.

2.5.1 Extraction of SVOC s / POM s from Samples

Extraction to gas phase, liquid phase and solid phase can be used in the prepara-tion of SVOC/POM samples. It is essential to estimate the recovery of each extrac-tion step. One method is spiking the sample with known amounts of internal standards similar to the analytes. The problem of this method is that spiked stan-dards may not bind to the matrix in the same way as the analytes. Another method is to fi nd the method with the highest recovery of a number of methods. Probably a combination of the two methods will give the most reliable results. Finally, use of certifi ed reference materials, if available, will be the best way to determine the total recovery.

Thermal desorption (TD) (see Section 2.6.1 ) compared with solvent desorption of adsorbent tubes requires less preparation since the samples are transferred on - line from the desorber to the GC. TD of indoor SVOC samples from Tenax TA tubes and quartz fi ber fi lters is possible (Clausen and Wolkoff, 1997b ) and air samples of phthalates from chamber studies are easily analyzed using TD (Clausen et al ., 2004 ). TD recovery may be estimated by comparing spiked tubes with direct solvent injection using the same GC (Clausen and Wolkoff, 1997b ).

When fi lter/adsorbent sampling is used no relevant information on the parti-tioning in the indoor environment can be obtained from separate extraction of the fi lter and the adsorbent. Therefore it is an advantage to extract the fi lter and the adsorbent together (Lewis, 1989 ; Sheldon et al ., 1993 ) and thus reducing the work-load by 50%. However, the extraction recovery of the fi lters and the adsorbent must be validated separately.

Comparison of simple methanol extraction, Soxhlet extraction, pressurized liquid extraction ( PLE ), and supercritical fl uid extraction ( SFE ) shows (Clausen et al ., 2003 ) that DEHP can be extracted relatively easily from dust and that the effectiveness does not differ signifi cantly between the different extraction methods (see Figure 2.4 ). Selection of the optimal method depends on several circum-stances, for example number of extraction cycles, instrument accessibility and the analysis method. However, PLE using cyclohexane/acetone was chosen as the preferred extraction method in the fi eld study.

2.5 Preparation of SVOC/POM Samples for Analysis 31

Preparation of commercial XAD - 2 glass sampling tubes is diffi cult since the tubes may contain both a sample section and a back - up section. The sections are diffi cult to remove and transfer to vials because the adsorbent grains are easily lost due to static electricity. To solve these problems a newly invented ‘ glass tube crusher ’ has been developed (Agner, 2005 ). When the glass tube has been crushed it can easily be transferred to an ASE extraction cell ( accelerated solvent extraction , Dionex) and be automatically extracted. The disadvantage of this method is that the back - up section is mixed with the sample section and two tubes in series (the second tube is back - up) must be used for sampling.

In order to avoid problems with thermal labile compounds during TD – GC analysis and to get a fast and gentle sampling and analysis technique a new air sampling ASE cell was developed by N ø rgaard et al . (2006) . The air sampling ASE cell has successfully been used to show the existence of ozonides in reaction mix-tures of ozone and limonene.

The most common method for the estimation of SVOC/POM content in mate-rial samples is solvent extraction of small pieces or fi nely ground powder of the material. The VOC content of PVC has been estimated in PVC powder which was obtained by liquid nitrogen cooling of the material before grinding (Cox, Little and Hodgson, 2001 ). Preferred extraction methods are PLE and SFE, since they are very effi cient. The materials must be extracted and the extracts analyzed several times to be able to estimate the number of extraction cycles needed for an exhaus-tive extraction.

2.5.1.1 Cleaning of Extraction Equipment Extraction equipment can be cleaned by extraction of blank samples using the same method as that for the real samples. For example, PLE cells used for phthal-ates (Clausen et al ., 2004 ) and PAHs (Jensen, Kofoed - S ø rensen and Clausen, 2005b ) have been cleaned including all accessories, fi lters, sand, etc., by a PLE method identical to the one used for extracting the real samples. Disassembled PLE extraction cells can be cleaned by sonication in acetone for at least 5 minutes. The purity of the extraction equipment is verifi ed by analysis of the blank extract and should be compared with the limits of detection.

Figure 2.4 Comparison of the extraction methods by the effectiveness of extracting DEHP from fl oor dust.


2.5.2 Concentrating Extracts of SVOC/POM Samples

The content of SVOCs in indoor air and dust samples is often relatively low com-pared with the limits of detection even on modern analytical instruments. Often it is necessary to sample over long periods or to concentrate the sample extracts before analysis. The most frequently used method is concentration by evaporation. However, there is a risk of losing SVOCs with high vapor pressure during the evaporation. A common method for concentrating the sample is to use a stream of an inert clean gas (e.g., nitrogen) while heating the sample to, for example, 40 ° C to gently evaporate the solvent to the desired volume. The gas is supplied from a small tube that is inserted in the sample vial close to the liquid surface of the extract. This concentration step can be very time - consuming and may not only evaporate the most volatile components (see Figure 2.5 ) but also the SVOCs of interest (Jakobsen et al ., 2003 ; Swartz, Stockburger and Gundel, 2003 ; Kofoed - S ø rensen and Clausen, 2004 ). Therefore, it is very important to validate the con-centration step before routine analysis. An additional pre - concentration step or a way to avoid the evaporation step is to use large volume injection (see Section 2.5.1 ) or solid phase extraction (see Section 2.6.1 ).

2.6 Analysis of SVOC s / POM s

The most widely used techniques for analysis of SVOCs and POMs are GC and HPLC. Those two techniques will be described here in relation to analysis of SVOCs/POMs. However, recently, very complex setups of the traditional tech-niques or quite new techniques have emerged. These techniques are used to obtain the analytical results easier and faster or to obtain even more information. An example is the automated analysis using direct thermal desorption of aerosol fi lter samples, two - dimensional GC and time - of - fl ight mass spectrometry (Vogt, Gr ö ger and Zimmermann, 2007 ). A recent innovation in mass spectrometry ( MS ) is the ability to record mass spectra on ordinary samples, in their native environment, without sample preparation by creating ions outside the MS instrument. In desorption electrospray ionization ( DESI ), electrically charged droplets are directed at the ambient object of interest; they release ions from the surface, which are then vacuumed through the air into a conventional mass spectrometer (Cooks et al ., 2006 ). DESI - MS has recently and for the fi rst time been used to directly analyze SVOCs in aerosol fi lter samples (Chen et al ., 2008 ).

2.6.1 Gas Chromatography ( GC )

GC is suitable for SVOCs/POMs that are stable at the temperatures used in GCs (up ca. 350 ° C) and can be eluted within an acceptable time ( < ca. 1 h). GC is a

2.6 Analysis of SVOCs/POMs 33

Figure 2.5 Recovery of PAHs using PLE with and without subsequent concentration by evaporation (Jensen, Kofoed - S ø rensen and Clausen, 2005b ) . The fi gure shows a loss of the most volatile PAHs by evaporation. The lower recovery of the higher PAHs which are not concentrated may be due to the low concentrations resulting in signifi cant loss to the walls of the vials.


relatively easy technique to use and has relatively low running expenses. GC offers high resolution that is important in analysis of SVOC/POM samples from indoor air since they can contain several hundred compounds (Clausen and Wolkoff, 1997b ). Furthermore, several different injection and detection techniques are available.

2.6.1.1 On - Column Injection ( OC ) On - column injection is preferred to splitless injection that exposes the injected compounds to high temperatures. In on - column injection the injection tempera-ture is usually about 10 ° C below the boiling point of the solvent in order to get the best chromatography. Thus, on - column injection is a gentle injection tech-nique suitable for less stable compounds.

2.6.1.2 Large Volume Injection ( LVI ) Large volume injection ( LVI ) is a very good alternative to pre - concentration of extracts by evaporation. The principle in LVI is to inject a large volume of sample extract on the column (50 μ l – 200 μ l) and venting out the large amounts of solvent before the real chromatography starts (see Figure 2.6 ). The method demands special equipment and testing time.

2.6.1.3 Thermal Desorption ( TD ) In TD the sample is transferred from the adsorbent tube by heating and collected in a cold trap ( − 30 ° ). Flash heating and purging of the trap carries the sample to the GC column. The advantage of TD compared to solvent extraction is its high

Figure 2.6 Large volume injection system.

2.6 Analysis of SVOCs/POMs 35

sensitivity, because the entire sample can be injected. However, while TD normally offers only one single analysis of the sample, solvent extraction gives the possibility to analyze the same sample several times.

Solvent extraction in combination with TD – GC – FID gives the opportunity to concentrate the extract directly on the adsorbent tube (solid phase extraction) by injection of the sample extract and purge off the solvent with for example, helium. Methanol extracts of house dust can be concentrated on adsorbent tubes by inject-ing up to 50 μ l and analyze the tubes by TD – GC (Kofoed - S ø rensen and Clausen, 2004 ).

The TD – GC system can be calibrated by running tubes that are spiked with a known volume and concentration of a standard vapor - gas atmosphere of the analyte or injecting a known concentration and volume of a solution of the analyte by a microliter syringe into the sorbent bed and remove the solvent by purging the tube with helium or nitrogen. The solvent must be tailored to the method, so that it can be evaporated from the tube without removing the analytes. Methanol or pentane are good choices for Tenax TA.

2.6.1.4 ‘ Cold Spots ’ and Other Adsorption Problems When working with SVOCs/POMs it is very important to ensure that there are no ‘ cold spots ’ , surfaces cooler than the intended fl ow - path temperature, coming into contact with the analyte.

At ‘ cold spots ’ the analytes can adsorb to the surface material and only a fraction will reach the detector. Subsequently, the adsorbed compounds can result in ‘ ghost peaks ’ in later analyses. Transfer lines from for example, a TD instrument to the GC must be insulated and heated to avoid ‘ cold spots ’ . The optimal temperature of the surfaces depends on the stability of the component, pressure, vacuum, etc. Some GC instruments have ceramic insulators at the inlet of the transfer line from for example, TD instruments and this material may act as a ‘ cold spot ’ if the line is not heated at the inlet.

2.6.1.5 Flame Ionization Detection ( FID ) Capillary GC in combination with a FID is an excellent and widely used technique to separate organic compounds. The advantage of the FID is the very wide linear dynamic range and the stability that retain a constant response for years if the fl ows remain stable. The samples must be desorbed either by solvent extraction or by TD.

2.6.1.6 Mass Spectrometric Detection ( MS ) A common method for identifi cation of organic compounds is mass spectrometry (MS) in combination with GC. After separation of the component by GC the mass spectrometer transform the analyte into gaseous ions in vacuum in the ion source. For electron impact ionization this results in different mass fragmentation pat-terns with different mass - to - charge ratios (m/z). From this fragmentation pattern it will be possible to identify the compound by comparison with commercial mass spectral libraries. Identifi cation of unknown compounds can be facilitated by


chemical ionization to obtain the molecular ion or by tandem MS (MS/MS) in which the ions can be fragmented into smaller pieces. Using GC – MS for quantita-tive determination usually requires internal standards, preferably deuterated target compounds, or stringent quality control such as for example, a control chart. The stability and the linear range of MS are not as good as for the FID detector, but MS has other advances for example, sensitivity, good ability for identifi cation and the possibility to check for interferences.

2.6.2 High Performance Liquid Chromatography ( HPLC )

HPLC is suitable for SVOCs/POMs that cannot be analyzed with GC due to their stability and/or low vapor pressure. HPLC is a more complicated technique than GC since there are many more parameters to vary (stationary phase, solvent, solvent compositions, solvent gradient, temperature etc.) and has relatively high running expenses due to the solvent use. A number of separation methods can be used in HPLC, for example ion exchange, gel fi ltration, normal phase and reversed phase. HPLC has lower resolution than GC but offers different separation tech-niques for special purposes as well as several detection techniques.

2.6.2.1 HPLC with Fluorescence Detection ( HPLC - FD ) Reversed phase is the most used separation technique and in combination with fl uorescence detection ( FD ) it is a good choice for analysis of for example, PAH (Hansen et al ., 1991 ) in indoor samples (Jensen, Kofoed - S ø rensen and Clausen, 2005a ). The FD is a relatively specifi c and sensitive detection method that uses both an excitation and an emission (fl uorescence) wave length and requires, of course, that the target compounds are fl uorescent. FD reduces the risk of interference.

2.6.2.2 HPLC with Mass Spectrometric Detection ( LC - MS ) LC - MS uses different types of soft chemical ionization that produces molecular ions and no fragmentation pattern. In MS/MS instruments the molecular ions can be fragmented by collision with a gas for example, He. This fragmentation can be used for identifi cation of a compound. No mass spectral libraries exist for LC - MS hence identifi cation of unknown compounds is more time - consuming than for GC – MS. For known compounds LC – MS is a very sensitive and specifi c method, using LC – MS/MS systems the analytical performance can be increased even more. LC - MS analysis is especially suitable for non - volatile POMs such as non - ionic surfactants in house dust samples (Clausen et al ., 2003 ).

2.6.3 Analysis Sequences

It is a good investment to have routines in building analysis sequences for quan-titative analyses on analytical instruments equipped with auto - samplers. It is

2.7 Quality Assurance and Control 37

always desirable to start the sequence with a check of the background of the instru-ment, the column and for example, the solvents by running blank samples. This could be followed by standards for the calibration. The position of the samples in the sequence should be randomized with controls samples (see Section 2.7.2 ), calibration standards, blank samples, and fi eld blanks distributed throughout the sequence. It should be considered to include a number of control samples (minimum two per sequence would be acceptable) in addition to an instrument control to check the response. The sequence could be fi nished with additional standards to check that the response remains stable throughout the analysis period.

2.7 Quality Assurance and Control

Development of a new sampling and analysis method can be a demanding job. In addition, it is time - consuming to build a complete quality control system. It is recommended to begin with some of the fundamental key points such as testing blank samples, making repeated runs, calibration curves, determining the limit of detection, and calibration of pumps, pipettes and balance. Also common sense and some experience are very valuable for generating good and reliable results. A full quality control system of a method makes it easier to conclude on results of a study, because the uncertainties are known and can be quantifi ed. Quality control can also be a gauge to help in trouble shooting (e.g., low response may indicate wrong split or leak). In SVOC analytics the demands on equipment and operators are very high and a well - designed quality control system will ensure a smooth operation and prevent misleading and inaccurate results.

2.7.1 Method Validation

Method validation is important to ensure that the analytical method is in statistical control. A method may be validated by the so - called method evaluation function ( MEF ) (Christensen et al ., 1993 ), which is obtained by linear regression analysis of the measured concentrations versus the ‘ true ’ concentrations. A ‘ true ’ concen-tration in a solution can be obtained by use of a high purity standard obtained from another manufacturer or batch than the one used for calibration. Both the high purity standard and the solvent are weighed using a traceable calibrated balance. If certifi ed reference material is available this is preferred. The method evaluation includes the most important characteristics of the method as the fol-lowing elements (see Figure 2.7 ):

• test for normal distribution of the residuals by the N - score test;

• test for linearity by the pure error lack of fi t test;

• test of slope different from one;


Figure 2.7 The method evaluation function (MEF) report for 2 - (2 - butoxyethoxy) ethanol (high conc. level) with plots of the ideal MEF, residuals and the total systematic error.

• test of intercept different from zero;

• estimation of the relative mean square error ( RMSE 1/2 ), which is a combined expression of the random and systematic errors;

• estimation of the limit of quantifi cation ( L Q ) and the limit of detection ( L D ). Both may be indicated by the upper tolerance interval that is shown in the RMSE 1/2 plot.


Validation must be carried out for a number of different components which refl ect a broad range of volatility and polarity or the target component. A series of minimum fi ve concentration levels with at least two samples of the analyte per level must be prepared. The MEF - samples must cover the concentration range of the method and should be run in random order on different days with fresh calibrations. The result of a MEF for TD – GC – FID of 2 - (2 - butoxyethoxy)ethanol (high level) (Hansen, Clausen and Wolkoff, 2001 ) is shown in Figure 2.7 . The MEF shows that the relationship between the measured results and the ‘ true ’ concentrations does not deviate from linearity. The distribution of the residuals shows no signifi cant deviation from normality. The slope (0.9929) does not differ signifi cantly from one, and the intercept (0.1618) is not signifi cantly different from zero. Thus, no systematic errors could be detected. All the used concentra-tions appeared to be above the L D and L Q since the plot of the RMSE ½ shows that the estimated RMSE ½ values are below the tolerance interval. The SD (0.2951) from this MEF is used in the internal Quality Control Chart. However, the accu-racy and precision of a method for one concentration level can be established by repeated measurement of traceable reference material. Precision is characterized as the standard deviation of the measurements and the accuracy as the difference between the mean of the measurements and the reference value, also called the bias.

2.7.1.1 Calibration Curves Calibration curves must be made from chemicals with the highest purity as pos-sible. To avoid dilution errors a multi - level calibration curve (six points) based on three stock solutions is recommended. One must also be aware that low concentra-tions of for example, PAHs (2 ppm) may be adsorbed by the vials up to ∼ 90% (Pinto, Jos é and Cordero, 1994 ). A calibrated and traceable balance or a traceable pipette must be used for accurate preparation and dilution of the standards. The calibration curve must cover the concentration range that is needed for the analy-sis. Both the slope and the intercept must be used to calculate the concentration in the sample, especially if the intercept is different from zero.

2.7.1.2 Limit of Detection ( L D ) and Limit of Quantifi cation ( L Q ) It is important to estimate the L D of the method, because results below the L D are encumbered with large uncertainty. L D is the lowest concentration of component in a sample which can be detected but not necessarily quantifi ed. It must be con-sidered whether results below the L D can be used or should be attributed the L D value, rejected or assigned a randomized value between zero and L D. The L D can be estimated by the MEF or estimated as three times the standard deviation of the analysis of 20 low standards. If it is decided to reject results below L D , L Q can be used as the lowest concentration of the component in a sample which can be quantifi ed. L Q can be estimated by the MEF or estimated as 10 times the standard deviation of the analysis of 20 low standards.

40

2 Sampling and A

nalysis of SVO

Cs and PO

Ms in Indoor A

ir

Figure 2.8 Example of a control chart with two control levels. The marked control sample indicates that the system may be out of control by use of the Westgard rules (Christensen et al ., 1993 ) .


2.7.2 Controls and Control Charts

To document that the analytical results are obtained by a method in statistical control an internal quality control system must be established. For routinely evalu-ation of the bias of the analytical system, a batch of stable control materials must be prepared. The control material should contain the target compounds or com-pounds with the same physical/chemical characteristic. In the control sample it could also be desirable to have other components that can give valuable informa-tion on the condition of the instrument, column, cold trap, splits etc. For checking response and retention time, the control analytes must be easy to analyze such as for example, alkanes. The control material should be stored under conditions at which it is stable and remain unchanged until analysis for example, in sealed ampoules or vials. There are different kinds of control chart systems. X - R and Westgard (see Figure 2.8 ) control chart are some of them. But the point is to construct a quality system which document that the accuracy and precision of the method stay within the decided and acceptable limits. The limits of the control chart can be estimated from the MEF, or determined by analysis of 10 – 20 control samples. The different control chart systems have different rules for rejection of runs that must not be violated. If one of the rules is violated it must be considered to reject the samples in the run. Samples analyzed in a sequence with rejected control components should not automatically be rejected since 5% of the controls statistically will be out of range. A summary of all the controls and all the control component analyzed on the same day indicates whether the analyses were out of statistical control or not. The internal control systems also includes analysis of blank samples, recovery testing, repetitive measurement of for example, certifi ed reference material and if possible participating in round robin tests such as Work-place Analysis Scheme for Profi ciency ( WASP ) program.

2.7.3 Documentation

A systematic and traceable documentation, for sampling, preparation and analysis of the samples, gives overview and saves time. Easy - to - use forms for the fi eld sampling with columns for sample numbers, pumps, date, sampling time, fl ow etc. should be prepared. A traceable number system must be build up, so the sample easily can be traced back to the sampling place. The notes must be legible and shortly describe the sample from sampling to results.


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References 43

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References 45

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Application of Diffusive Samplers Derrick Crump

3

3.1 Introduction

A diffusive/passive sampler is a device which is capable of taking samples of gas or vapor pollutants from the atmosphere at a rate controlled by a physical process such as diffusion through a static air layer or permeation through a membrane, but which does not involve the active movement of air through the sampler (Brown, 1993 ). Compared with methods requiring pumps for active air movement, diffusive samplers are more convenient and have lower associated sampling costs. They are particularly useful for determining long - term mean concentrations of pollutants in ambient and indoor air over periods typically of one day to four weeks.

The development and evaluation of the performance of diffusive samplers in the 1980s and early 1990s was driven by requirements for workplace applications. A European standard (EN 838:1995) was published describing tests to assess effects of analyte concentration, environmental conditions and exposure time on sampler performance as well as accuracy and reproducibility (EN, 1995 ). A number of commercially available samplers were evaluated for workplace monitoring. The UK Health and Safety Executive (HSE, 1995 ) published recommended methods for the determination of VOCs based on sampling with a tube type sampler con-taining a sorbent and analysis by TD – GC, and a method for the determination of hexane in air using badge type samplers which are solvent desorbed and the eluent is analyzed by GC (HSE, 1992 ).

Since the mid - 1990s the use of diffusive sampling for the measurement of VOCs in indoor and ambient air has increased markedly. Experience gained from devel-oping methods appropriate for the workplace has formed a basis for a number of studies leading to development of methods appropriate for environmental applica-tions, including indoor air. Such was the knowledge gained, that by 2003 an international standard was agreed providing general guidance on the diffusive sampling and analysis of VOCs in indoor, ambient and workplace air (EN ISO, 2003 ). A further specifi c standard for the measurement of formaldehyde in indoor


47

48 3 Application of Diffusive Samplers

air by use of a diffusive sampler was published in 2004 (ISO, 2004c ) as well as a European standard on guidance for selection, use and maintenance of diffusive samplers for determining gases and vapors (including VOCs) in indoor air (EN, 2004b ).

These international standards are indicative of the recognition that diffusive samplers are an important tool alongside other methods such as short term pumped sampling and continuous monitoring for the study of VOCs in indoor environments. The appropriateness of particular methods depends upon the objectives of the measurement and availability of methods. An international stan-dard on sampling strategies for indoor air provides guidance on selecting appropri-ate methods (ISO, 2004a ) and further standards give strategic guidance specifi c to VOCs and formaldehyde (ISO/FDIS, 2007 , ISO, 2004b ).

This chapter discusses the principles governing diffusive sampling and the factors that can infl uence sampler performance, and reviews some of the studies that have applied the technique for the measurement of VOCs in indoor air.

3.2 Principles of Diffusive Sampling

In the ideal case the resistance to mass transfer of a diffusive sampler is confi ned to the stagnant air gap between the sampling face of the device and the surface of the collecting sorbent material (Van den Hoed and Van Asselen, 1991 ). Then the mass fl ow through the sampler can be described by Fick ’ s fi rst law of diffusion

J DA c x= − d d (3.1)

J = diffusion fl ux (g/s), D = air diffusion coeffi cient (cm 2 s − 1 ), A = cross sectional area of the sampler (cm 2 ) and d c /d x = the concentration gradient of the compound across the stagnant air gap.

Assuming the compound is effectively trapped by the sampler, the concentration at the interface of the air gap and collecting surface ( C e) will be zero. Then the mass collected by the sampler ( M ) is given by

M Kt C= ( )e o (3.2)

where t (e) = the exposure time, C o = the concentration of the compound in ambient air and K = the uptake rate of the sampler defi ned by

K DA L= (3.3)

where L = the length of the stagnant air gap (cm). The expression DA / L has units of cm 3 s − 1 and represents the diffusive uptake

rate of the sampler under ideal conditions.

3.2 Principles of Diffusive Sampling 49

Equation 3.2 assumes a zero concentration at the interface of the air gap and collecting surface. Samplers that rely on chemisorption or other reactive trapping will maintain a zero concentration providing the capacity of the trapping agent is not exceeded. However if the trapping is based on reversible sorption, the ‘ perfect sink ’ situation at the collecting surface may not be achieved and the effective sampling rate will change during the sampling period. In addition sample loss, or reverse diffusion, will occur if the ambient concentration decreases below that which is in equilibrium with the sorbent surface.

Equation 3.3 enables the ideal uptake rate of the diffusive sampler to be cal-culated with knowledge of A , L and the diffusion coeffi cient for the analyte of interest. Mass transfer considerations limit the maximum sampling rate that can be achieved for a given air velocity in the environment being monitored. If the environment at the face of the sampler cannot be replenished with more of the analyte of interest at a rate greater than that demanded by the sampler then the uptake rate will be lower than that predicted for the ideal case. This can be described as ‘ starvation ’ of the sampler and is of particular concern when under-taking stationery monitoring in indoor environments with low air speeds. Con-versely high air velocities can increase the sampling rate by causing turbulence inside the air gap of the sampler and thus reduce the effective diffusive path length of the sampler.

For indoor environments the effects of changes in temperature and pressure on diffusive uptake rate will be insignifi cant compared with other sources of error. High humidity can affect the sorption capacity of hydrophilic sorbents such as charcoal. This will reduce the time before saturation of the sorbent occurs.

If the concentration at the face of the sampler fl uctuates there is the possibility that the concentration inside the sampler air gap will not refl ect the actual ambient concentration. The sampler would then not provide a true integrated response to the exposure concentrations. The response time of passive samplers is typically 1 – 10 seconds and provided that the total sampling time is large relative to the response time, errors will be small (Brown, 1993 ).

It is necessary to recover the analyte from the sorbent by solvent extraction or thermal desorption. A weaker sorbent allows more effi cient recovery during analy-ses but will not act as the ‘ perfect sink ’ at the collecting surface. No single sampling or analytical technique is suitable for measuring the wide range of compounds that occur in indoor air.

This consideration of the principles of diffusive sampling identifi es a range of factors which may infl uence the performance of a diffusive sampler for monitor-ing VOC concentrations in indoor air. These factors will potentially be a source of error in such measurements and add to the overall uncertainty of the result given by the measurement procedure. In addition the amount of uncertainty will be infl uenced by other factors including amount and consistency of background contamination of sorbents, repeatability of analytical determination, formation of artifacts, stability of analyte on the sorbent, recovery of analyte during analyses and presence of interferents.


3.3 Selection of Appropriate Methods

As well as selecting methods with acceptable performance in terms of accuracy and precision, the sampling method should be in accordance with a strategy that ensures a representative measurement of the environment of interest. Guidance on designing a sampling strategy to meet different sampling objectives in the indoor environment has been published by several groups (ECA, 1994 , ISO, 2004a , EN, 2004b , Crump et al ., 2002 ). These objectives could include the following:

• identifi cation of causes of complaints about indoor air quality; • determining reference values for particular pollutants to establish knowledge of

normal and abnormal levels of pollutants; • checking for compliance with air quality standards and guidelines; • identifi cation of pollution sources; • testing the effectiveness of remedial actions aimed at reducing pollution; • determination of trends in air quality; • determining personal exposure to pollution; • validation of indoor air pollution models.

The selection of appropriate methods will depend upon the pollutant of interest, the averaging time of any guideline or standard, the availability of methods, the duration of measurement, the sensitivity required and resources available. Dif-fusive samplers can be particularly suited to the measurement of a compound such as benzene where the health concern relates to long - term exposure and air quality standards are based on annual mean concentrations. Diffusive samplers may be less applicable for determining exposure during particular activities that may occur for a period of minutes or hours (e.g., cooking, painting), or if checking for compliance with a guideline value such as formaldehyde based on a 30 minute exposure period.

3.4 Performance of Diffusive Samplers for the Measurement of VOC s in Indoor Air

Most users of diffusive samplers are unlikely to be able to undertake a full evalu-ation of sampler performance and will look to manufacturers and published studies for evidence that the sampler is fi t for purpose. This knowledge is improv-ing rapidly, but the user should consider available knowledge to identify limita-tions and whether this will impact on the uncertainty of measurement for their own application. Even in the case where good data are available, it is necessary for a laboratory to undertake their own assessment of precision and accuracy and where possible ensure full traceability to primary standards which is a requirement for third party method accreditation.

For workplace monitoring two main types of diffusive sampler have been used for monitoring of indoor air:

3.4 Performance of Diffusive Samplers for the Measurement of VOCs in Indoor Air 51

i) Badge type samplers containing a strong adsorbent such as charcoal or a reac-tive chemical, for example DNPH for carbonyls, that requires solvent desorp-tion for GC or HPLC analysis. An example is the OVM 3500, a circular badge with a 1 cm diffusion length containing a charcoal wafer. Desorption of VOCs is carried out within the monitor itself by the addition of carbon disulphide, a toxic solvent that requires appropriate disposal. The recovery of some com-pounds can be poor and contaminants in the solvent can reduce sensitivity as does the dilution effect; typically 2 ml of solvent is used to desorb and 1 μ l is used for GC analysis (i.e., 0.05% of the collected mass of analyte). Exposure periods applied have ranged from 24 hours to several weeks. The diffusive uptake rates reported by the manufacturer for 8 - hour exposure periods are about 30 ml min − 1 but actual values are compound specifi c.

ii) Tube type samplers with weaker adsorbents such as the porous polymer Tenax and carbonaceous powders that can be thermally desorbed. Most widely used has been the Perkin Elmer type sampler which was specifi cally designed for TD. It consists of a 90 × 6.3 mm OD (5 mm ID) steel tube within which an adsorbent is retained by a stainless steel mesh. The rear end of the tube is sealed with a brass Swagelock fi tting and PTFE ferrule. At the sampling end there is a 1.53 cm air gap between the mesh containing the adsorbent and the end of the tube. Sampling rates are of the order of 0.5 ml min − 1 .

A diffusive end cap which contains a stainless steel mesh screen can also be used and this can prevent air movement within the diffusive air gap when exposed to high air velocities (Brown, 1993 ). Evaluation of the sampler for workplace monitor-ing found that the diffusive uptake rate of the sampler was not infl uenced by air velocities as low as 0.007 m s − 1 .

Other types of diffusive sampler have been less widely applied for indoor air studies. These include tube type samplers that are solvent desorbed and radial type samplers consisting of a cylindrical adsorbent surface that has a short diffusive path resulting in an effective uptake rate that is typically 100 times that of the tube type sampler (Cocheo, Boaretto and Sacco, 1996 ). One type of radial sampler developed for measuring carbonyl compounds in indoor air comprises silica gel coated with 2,4 - DNPH as the adsorbent within a sintered polyethylene tube that acts as a diffusive membrane (Uchiyamaa, Aoyagi and Ando, 2004 ).

The tube type samplers have lower diffusive uptake rates and are therefore less prone to ‘ starvation ’ effects. In consequence they require longer exposure periods to collect the same mass of analyte. The occurrence of reverse diffusion depends on the analyte - sorbent interaction, but for weaker adsorbents such as Tenax TA this does occur with more volatile compounds for example, compounds more volatile than toluene.

Back diffusion can be prevented by use of stronger adsorbents such as graphi-tized carbon but recovery of less volatile compounds by thermal desorption is less effi cient. The analysis can be automated and commercial equipment is available. The samplers can be re - used many times without loss of performance. Sensitivity of TD – GC is potentially higher because typically 20 – 30% of the collected analyte


is transferred to the GC column although this has an associated disadvantage because each sampler can only be analyzed once. However some commercially available equipment enables the re - collection of desorbed analyte that is not trans-ferred to the chromatography column and this can subsequently be analyzed quantitatively.

The fi rst protocols developed for evaluation of the performance of diffusive samplers were based on workplace applications. The European standard EN838:1995 (EN (1995) ) is an example. This approach has been adapted to provide a protocol for the evaluation of the performance of diffusive samplers for ambient air monitoring (EN, 2004a ). It describes a series of tests that enable a calculation of the measurement uncertainty. The key sampler related factors assessed are:

• desorption effi ciency (using either solvent or thermal desorption); • bias due to selection of a non - ideal sorbent (for samplers relying on reversible

sorption); • effect of air velocity/sampler orientation; • effect of environmental parameters (temperature and humidity) on sampler

performance; • shelf life prior to use; • sampler leak test; • sampler integrity after exposure and prior to analysis; • blank value; • fi eld tests under conditions representative of practical extremes to be encountered

during the intended application.

An example of an evaluation undertaken according to this protocol is for the mea-surement of benzene in ambient air by tube type samplers subject to thermal desorption (EN, 2005b ) and samplers subject to solvent desorption (EN, 2005a ). These standards were developed specifi cally to provide measurement methods meeting requirements for checking compliance with the air quality standard for benzene set by the European Ambient Air Quality Directive ( Directive 96/62/EC, 1996 ). Evaluation of published data concerning the samplers under the test condi-tions enabled an evaluation of the expanded relative uncertainty of measurement at a limit value (5 μ g m − 3 ) for benzene in air and for solvent and thermally desorbed samplers a value of ± 13.4% was obtained.

The method of measuring benzene in air can be considered suitable for indoor air monitoring although particular consideration should be given to the impact of low air velocities that may be present in indoor environments when selecting an appropriate sampler. Generally information for determination of other VOCs is less comprehensive. A possible exception is formaldehyde using a badge type sampler consisting of a fi lter coated with DNPH which is described in ISO 16000 - 4:2004 (ISO (2004c) ) along with performance data. However, a full overall uncer-tainty evaluation is not included. The standard EN ISO 16017 - 2:2003 (EN ISO (2003) ) summarizes diffusive uptake rates appropriate for ambient and indoor air applications for sorbent tubes reported in the literature for benzene, toluene, xylene, ethylbenzene, trimethylbenzene and decane, as well as data on recovery by

3.5 Studies of VOCs in Indoor Air Using Diffusive Samplers 53

Table 3.1 Ideal and experimentally determined uptake rates (ng ppm min − 1 ) for chlorinated butadienes and hexachloroethane using Tenax TA and a 4 week exposure period (Hafkenscheid et al ., 2001 ).

Compound U id U Eeff CV% for U Eeff

Hexachloroethane 4.7 2.4 7.0 Tetrachlorobutadiene 3.9 3.0 1.7 Pentachlorobutadiene 4.3 3.4 2.2 Hexachlorobutadiene 4.6 3.5 1.7

thermal desorption and stability during storage. Some additional information can be found in the scientifi c literature, an example being the determination of 4 - week diffusive uptake rates for the determination of chlorinated butadienes and hexa-chloroethane in indoor air (Hafkenscheid et al ., 2001 ). Laboratory tests using standard atmospheres were used to determine the effective diffusive uptake rates of sorbent tubes containing Tenax TA that were analyzed by TD – GC. The results of rates experimentally determined ( U eff ) and the ideal rates ( U id ) derived by cal-culation are summarized in Table 3.1 . The differences between the ideal and measured uptake rates is notable and to be expected from knowledge of the non - ideal behavior of sorbents.

In the absence of experimental data for diffusive uptake rates it is possible to calculate the diffusive uptake rate based on fundamental principles. Standards EN ISO 16017 - 2:2003 (EN ISO (2003) ) and EN 14412:2004 (EN (2004b) ) provide guid-ance on undertaking these calculations. However as illustrated by Hafkenscheid et al . (2001) the actual behavior may not be ideal. One consequence is that the effective uptake rate may change with length of exposure period. Diffusive uptake rates for six VOCs were determined experimentally by Brown, Crump and Yu (1993) who exposed Perkin Elmer type tubes packed with Tenax TA in standard atmospheres for periods of 7 to 42 days. The results showed a signifi cant decline in the effective sampling rate with exposure time for the more volatile compounds (e.g., benzene and toluene) and no signifi cant change for compounds less volatile than xylenes (see Figure 3.1 ).

3.5 Studies of VOC s in Indoor Air Using Diffusive Samplers

Diffusive samplers have been applied in major studies of indoor air quality includ-ing national surveys, particularly in homes. The benefi ts of relatively low cost, no noise and unobtrusive appearance, the practicality of placement by persons in receipt of a relatively low level of training and ability to mail samplers between the laboratory and fi eld site are important for these applications.


A national survey of indoor air quality in England included the measurement of formaldehyde and VOC concentrations using diffusive samplers in 876 homes (Raw et al ., 2004 ). The study was designed to increase knowledge of baseline pol-lutant levels and factors associated with high concentrations. Formaldehyde levels were measured over a period of 3 days in the bedroom of each home using a passive badge type sampler that consists of a fi lter paper impregnated with DNPH that is retained in a plastic housing. VOCs were determined by diffusive sampling tubes (Perkin Elmer type) packed with Tenax TA adsorbent with an exposure period of 4 weeks. The bedroom was selected for sampling as it is the room in which most people spend the largest amount of time indoors. Total VOC ( TVOC ) concentrations were determined as well as the concentration of 22 individual VOCs. Table 3.2 summarizes the results. In about 5% of homes the TVOC con-centration exceeded 1000 μ g m − 3 and the median value for all homes was 200 μ g m − 3 . Formaldehyde levels exceeded 100 μ g m − 3 in 0.7% of the homes. There were sea-sonal differences in the TVOC concentration with highest mean concentrations occurring in autumn. Concentrations were higher where painting had occurred in the home during the sampling period or the previous 4 weeks, and homes with an integral garage were higher than those with a detached garage or no garage. After adjusting for homes that had painting undertaken, concentrations of TVOC were higher in newer homes and in bedsits and fl ats than other types of dwellings. Formaldehyde varied signifi cantly with building age, newer homes having higher concentrations. The presence of particleboard fl ooring was associated with higher formaldehyde concentrations.

Table 3.3 summarizes results of other major studies that employed the use of diffusive samplers to measure VOCs and formaldehyde in indoor air in UK homes.

Figure 3.1 Experimentally determined diffusive uptake rates for toluene using different exposure periods and the Perkin - Elmer type sampler packed with Tenax TA.


Table 3.2 Concentrations of VOC including formaldehyde ( μ g m − 3 ) measured in major indoor air quality studies using diffusive samplers (Raw et al ., 2004 ; Rehwagen, Schlink and Herbarth, 2003 ; Kirchner et al ., 2003 ; Jia et al ., 2005 (1); Otson, Fellin and Tran, 1994 ); Cohen et al ., 1989 (2)).

Compound England Germany France USA (1) USA (2) Canada

G M Max A M Max G M A M A M A M

Formaldehyde 22.2 171 – – 20.8 – – – TVOC 210 3360 200 4597 – 187 – – Benzene 3.0 94 3.1 41 2.1 4.1 6.7 5 Toluene 15.1 1784 31.8 814 15.8 28 – 41 m+p xylene 3.8 153 9.6 2496 5.1 9.8 17.7 20 Limonene 6.2 308 36.7 1278 12.9 17.6 – 20 Undecane 2.6 247 10.2 582 7.6 – – – TPDMIB 5.1 770 – – – – – – TPDDIB 1.6 100 – – – – – – Heptane 2.9 181 8.4 328 – – – – Methylcyclo hexane 4.9 411 5.3 172 – 1.5 – – Octane 1.5 90 2.9 480 – – – – Hexanal 0.9 56 – – – – – – Nonane 3.8 213 4.3 169 – 6.9 – – Ethyl benzene 1.2 54 3.6 386 2.1 – 5.4 8 α - pinene 5.0 306 24.8 393 5.1 21.3 – 23 Decane 4.5 481 10.0 808 7.4 – 20.6 31 3 - carene 1.6 94 8.9 382 – – – –

1,2,4 - trimethylbenzene 1.9 430 – – 2.9 3.7 11.5 12 Benzaldehyde 1.1 84 – – 0.8 – – – DMCPS 5.9 982 – – – – – – Nonanal 3.2 57 – – – – – – Tetradecane 0.8 13 – – – – – – Hexadecane 0.7 12 – – – 0.8 – –

DMCPS = decamethylcyclopentasiloxane, TPDMIB = 2,2,4 - trimethyl - 1,3 - pentanediol monoisobutyrate, TPDDIB = 2,2,4 - trimethyl - 1,3 - pentanediol diisobutyrate, G M = geometric mean, A M = arithmetic mean.

The BRE Indoor Environment study monitored concentrations of formaldehyde and VOCs (as well as nitrogen dioxide and biological particulates) in 174 homes in Avon (South West England) over a 12 month period (Berry et al ., 1996 ). The study was conducted between 1990 and 1993. All participants were expectant women when they joined the study and they were enrolled in the ALSPAC study of childhood health. The methods were essentially the same as those used in the Indoor Air Quality Survey of England. Some of the main fi ndings were that form-aldehyde concentrations were higher in newer homes, TVOC concentrations were highest in homes recently decorated and benzene concentrations were higher in homes of smokers and those with integral garages in which a car was kept.


Table 3.3 Mean (arithmetic) concentrations of air pollutants in UK buildings ( μ g m − 3 ).

Study TVOC Benzene Formaldehyde

In Out In Out In Out

174 homes in Avon (Berry et al ., 1996 ) 415 40 8.0 5 25 2

40 homes in Southampton (Brown and Crump, 1998 )

330 57 6.0 6 26 2

100 home and offi ce buildings in UK where concern about IAQ ( Brown, Cockram, Crump, Mann, 1 996)

1040 – – – 110

30 homes and offi ces in Hertfordshire (Crump et al ., 1999 )

388 77 6.1 3.9 29 4 167 a 3.0 a 24 a

500 homes in Nottingham (Venn et al ., 2003 )

448 – 6.0 – 25 –

10 homes and 10 offi ces in Southern England (Mohle et al ., 2003 )

200 9 2.6 1.3 28 – 86 a 28 a 1.4 a 2.0 a 25 a –

37 homes (winter) in Southern England built since 1995 (Dimitroulopoulou et al ., 2005 )

247 17 1.9 1.4 30 3

a Offi ces.

Other major UK studies using similar diffusive sampling methods are; the effect of intervention measures on the incidence of symptoms of asthma involving 40 patients in Southampton (Brown and Crump, 1998 ), the relationship between concentrations in indoor and outdoor environments and personal exposure (Crump et al ., 1999 ), air quality in 400 homes of asthmatic and non - asthmatic children in Nottingham (Venn et al ., 2003 ), the measurement of air pollutants in buildings with occupants concerned about poor indoor air quality (Brown et al ., 1996 ), validation of sampling methods in 10 homes and offi ces (Mohle et al ., 2003 ) and the study of indoor air quality and rates of ventilation in homes in England built since 1995 (Dimitroulopoulou et al ., 2005 ). Crump et al . (2007) describe the use of tube type diffusive samplers containing the sorbent Carbograph ITD to determine mean four week concentrations of benzene in indoor and outdoor air in a study of 54 homes in Scotland. The purpose was to investigate the possible ingress of benzene into buildings from ground contaminated by the historic use of the site for the manufacture of gas from coal. Monitoring was undertaken over


a 12 month period and no signifi cant effect of ingress on indoor benzene concen-trations was found, but homes with an occupant who smoked tobacco had higher concentrations in the indoor air than homes with no smoker.

Further specifi c studies have used diffusive samplers to investigate the impact of emissions from cars kept in an integral garage on benzene concentrations in the indoor air (Mann, Crump and Brown, 2001 ) and to assess the effectiveness of remedial measures used to reduce radon ingress into homes for controlling ingress of chlorinated butadienes present in contaminated land (Crump et al ., 2004 ).

The mean concentrations of pollutants in Table 3.3 do not differ greatly between the studies. The indoor/outdoor ratios refl ect the dominance of indoor sources for TVOC and formaldehyde and the greater importance of outdoor air for benzene (where the main source is traffi c) for determining the indoor concentration. The measurements of formaldehyde and TVOCs undertaken in homes with occupants concerned about the air quality were considerably higher than mean concentra-tions in the other studies.

In France, a permanent national survey of indoor air quality has been estab-lished (Kirchner et al ., 2003 ). A pilot investigation was conducted in 2001 involving 90 dwellings across three geographical areas. Radial passive samplers were used, one type containing a Carbograph cylinder that could be thermally desorbed for VOCs and another impregnated with DNPH for aldehydes that was solvent desorbed. Results for selected VOCs based on 62 measurements in kitchens are included in Table 3.2 .

In Germany diffusive samplers have been applied extensively to monitor VOCs in indoor air. Rehwagen, Schlink and Herbarth (2003) used data generated between 1994 and 2001 within several epidemiological studies. They summarize results of 1499 indoor (apartment) and 222 outdoor measurements in Leipzig using the badge type OVM 3500 sampler prior to 1999, and the Perkin Elmer type tube thereafter, with a 4 week exposure period. The VOC load indoors was about ten times that outdoors and winter concentrations were higher than summer. They noted a downward trend in VOC concentrations over the period 1996 to 1999, except for terpenes which increased. The arithmetic mean and maximum concen-trations of some VOCs are shown in Table 3.2 .

The Perkin Elmer type tube packed with Tenax was used in a study of exposure to air pollutants of 100 offi ce workers in Milan, Italy (Cavallo et al ., 1997 ). Details of exposure time and analytical methods are not given in the paper. Concentrations of TVOC in homes (mean 411 μ g m − 3 ) as well as toluene and xylene concentrations were very similar to those measured in homes in other European studies although benzene concentrations (mean 28.8 μ g m − 3 ) were considerably higher.

In Sweden thermal desorption tubes containing Tenax TA were used to measure VOCs in 178 randomly selected dwellings using a 30 day exposure period (Borne-hag and Stridh, 2000 ). More than 100 VOCs were identifi ed with a mean concen-tration less than 25 μ g m − 3 . Temperature and ventilation rate were not correlated with VOC concentrations, but there was a correlation with relative humidity to some extent.


The OVM 3500 monitor was applied by Cohen et al . (1989) to determine VOC concentrations in a study of 35 homes in West Virginia, USA. Samplers were placed in the main bedroom at a height of approximately 1.5 m, and outdoors in a specially fabricated aluminum shelter that kept the sampler dry. They were exposed for one 3 week period before analysis by solvent desorption using 1.5 ml carbon disulphide and analysis with GC – FID. Results for indoor concentrations of some of the 17 compounds reported are shown in Table 3.2 .

Jia et al . (2005) undertook measurements in 30 – 50 homes in each of 3 cities in Michigan, USA using a tube type sampler that was thermally desorbed and an exposure period of 3 – 4 days. 49 compounds were identifi ed indoors and indoor/outdoor ratios ranged from 1 to 82. Signifi cantly higher benzene, toluene and naphthalene concentrations occurred in homes with attached garages and D - limonene was higher in homes with smokers. Concentrations in air reported in one city are summarized in Table 3.2 .

A nationwide study of indoor air concentrations of 26 VOCs was conducted in Canada in 1991 (Otson, Fellin and Tran, 1994 ). An OVM - 3500 sampler was exposed for 24 h on one occasion in 757 homes. Approximately equal numbers of homes were sampled in each of four seasons. VOCs were extracted with 1.5 ml of carbon disulphide and 26 compounds were determined by GC with a mass selec-tive detector operated in the selected ion monitoring mode. Detection limits were estimated to be in the range 1.6 – 5.9 μ g/m 3 . Mean concentrations of selected com-pounds are shown in Table 3.2 .

In Australia badge type samplers have been used to measure formaldehyde concentrations in more than 100 caravans (Dingle, Tan and Cheong, 2000 ) and also in homes of 224 children. In the homes study formaldehyde concentrations greater than 50 ppb were associated with an infl ammatory response in children ’ s lungs (Franklin, Dingle and Dick, 2000 ). Formaldehyde has also been the subject of considerable study in Japan including a study by Haruki et al . (2003) of 300 government buildings using a 24 - hour exposure period where 1.5% of readings exceeded a guideline value of 0.08 ppm.

A number of studies included the use of diffusive samplers to measure personal exposure by attaching samplers to the person as well as the monitoring of fi xed site locations. Ni et al . (2005) used a tube type sampler containing charcoal that was solvent desorbed to measure VOCs and a cartridge containing DNPH to measure formaldehyde in 160 homes in China. The mean indoor concentration of benzene was 27.9 μ g m − 3. This was higher than the outdoor concentration (19.6 μ g m − 3 ), but similar to the measured personal exposure (27.7 μ g m − 3 ). Indoor formaldehyde concentrations (mean 64.7 μ g m − 3 ) exceeded those measured out-doors and were similar to personal exposure concentrations. Modig et al . (2004) used tube type samplers containing the adsorbent Carbopack X with a one week exposure period to monitor 40 persons in Sweden and these were thermally desorbed. The average benzene concentration was 2.2 μ g m − 3 .

Ullrich et al . (1996) used the OVM 3500 sampler in a study of indoor, outdoor and personal exposure concentrations involving 156 people in eastern Germany. The diffusive sampler was worn by subjects for either 2 days or 2 weeks. Samplers were also exposed simultaneously to measure VOCs in indoor and outdoor air.

3.7 Conclusion 59

The authors compared the results with the previous measurements undertaken in homes in western Germany and concluded that there were no remarkable differ-ences between the two groups studied. The indoor air results were broadly similar to the personal exposure measurements which for most VOCs were considerably higher than for the outdoor air. Hoffman et al . (2000) reports the use of diffusive badges in the German Environment Survey in 1990/92 where some 70 VOCs were determined. Benzene concentrations correlated with exposure to ETS, and paint-ing and smoking correlated with amount the of C 9 aromatics. Crump et al . (1999) studied the personal exposure to VOCs and formaldehyde of 30 adults in England and found these to be higher than indoor concentrations recorded by indoor moni-tors in homes and workplaces. Indoor concentrations were higher than outdoor concentrations.

3.6 Other Applications of Diffusive Samplers

An application of growing interest internationally is the measurement of the air exchange rate of rooms and buildings with outdoor air by the release of a VOC used as a tracer gas and the measurement of its concentration using a diffusive sampler. Examples are studies by Stymme, Emenius and Boman (2005) who used passive tracer gas measurements to measure air exchange with the outdoors in two homes in Sweden for 47 two week periods and Dimitroulopoulou et al . (2005) who report the measurement of air exchange in 37 homes in England over 2 week periods in summer and winter. Both studies involved placement of sources of a perfl uorocarbon compound in all rooms and the measurement of the mean con-centration of tracer in the air using adsorbent tube samplers placed in each room that were analyzed by TD – GC. The rate of air exchange is determined by mass balance calculations as the amount of tracer released is known as well as the room volumes and concentration of tracer in air.

Diffusive samplers have also been developed to determine SVOCs but there have been relatively few studies to date. An example is the passive fl ux sampler devel-oped by Fujii et al . (2003) to determine the rate of emission of phthalate esters from materials. The sampler consisted of a circular metal disc containing activated carbon particles held within an inert matrix of PTFE. The sampler was placed on the material under test giving a diffusion length of 0.5 or 2 mm depending upon the design and adsorbed phthalate esters were extracted from the sampler with toluene and determined by GC – MS.

3.7 Conclusion

Diffusive samplers originally developed for workplace monitoring have been applied to a wide range of studies of VOCs in non - industrial indoor air through a modifi cation of sampling times and analytical methods. The main types of


samplers used for major studies of VOCs in indoor air are badge type which are usually solvent desorbed prior to analysis by gas chromatography and the tube type designed for thermal desorption. In addition radial samplers with high rates of sampling have been used in a few studies.

Guidance on the selection of methods and their performance are available in some European and International standards. There is a need for further validation and assessment of the overall uncertainty of measurement by diffusive methods although this data is increasingly available, as for example with standard methods for determining benzene. The techniques are particularly suited to the determina-tion of mean concentrations of VOCs over periods of a day to several weeks for both fi xed site and personal monitoring. The investigator needs to consider care-fully the choice of sampler depending upon the objective of the study, the analyte of interest and the available information on sampler performance. Considerations include (i) the possibility of ‘ starvation ’ during sampling which will depend upon the type of sampler employed and the air movement in the environment to be monitored, (ii) the detection limit required which may be limited by the amount and consistency of contaminants in unexposed samplers and (iii) whether automa-tion of procedures and avoidance of solvents are important factors.

Diffusive samplers have proved to be the preferred approach for large scale surveys of indoor air quality and are being increasingly used for non - occupational personal exposure studies and the measurement of rates of ventilation. The low costs and convenience compared with pumped methods ensure a growing use of the technique for health studies, building investigations and the study of trends in indoor air quality that result from changes in outdoor air quality, construction practice, building operation and occupant behavior.

References

Berry , R. , Brown , V. , Coward , S. , Crump , D. , Gavin , M. , Grimes , C. , Higham , D. , Hull , A. , Hunter , C. , Jeffery , I. , Lea , R. , Llewellyn , J. and Raw , G. ( 1996 ) Indoor air quality in homes: the Building Research Establishment, Indoor Environment Study . BRE Report BR 299 and BR 300 , CRC Ltd , Watford, UK .

Bornehag , C. - G. and Stridh , G. ( 2000 ) Volatile organic compounds (VOC) in the Swedish housing stock . Proceedings of Healthy Buildings 2000 , August 6 – 10, Espoo, Finland , Vol. 1, pp. 437 – 42 .

Brown , R. ( 1993 ) The use of diffusive samplers for monitoring of ambient air . Pure and Applied Chemistry , 65 ( 8 ), 1859 – 74 .

Brown , V. and Crump , D. ( 1998 ) The use of diffusive samplers for the measurement of

VOCs in the indoor air of 44 homes in Southampton . Indoor and Built Environment , 7 , 245 – 53 .

Brown , V. , Crump , D. and Yu , C. ( 1993 ) Long term diffusive sampling of volatile organic compounds in indoor air . Environmental Technology , 14 , 771 – 7 .

Brown , V. , Cockram , A. , Crump , D. and Mann , H. ( 1996 ) The use of diffusive samplers to investigate occupant complaints about poor indoor air quality . Proceedings of Indoor Air ’ 96 , July 21 – 26, Nagoya, Japan , Vol. 2, pp. 115 – 20 .

Cocheo , V. , Boaretto , C. and Sacco , P. ( 1996 ) High uptake rate radial diffusive sampler suitable for both solvent and thermal desorption . American Industrial Hygiene Association Journal , 57 , 897 – 904 .

References 61

Cavallo , D. , Alcini , D. , Carrer , P. , Basso , A. , Bollini , D. , Lovato , L. , Vercelli , F. , Visigalli , F. and Marconi , M. ( 1997 ) Exposure to air pollutants in homes of subjects living in Milan . Proceedings of Healthy Buildings 1997 , Washington DC, USA, September 27 – October 2, 1997 , Vol. 3, pp. 141 – 5 .

Cohen , M. , Ryan , P. , Yanagisawa , Y. , Spengler , J. , Ozkaynak , H. and Epstein , P. ( 1989 ) Indoor/outdoor measurements of volatile organic compounds in the Kanawha Valley of West Virginia . Journal of the Air Pollution Control Association , 39 , 1086 – 93 .

Crump , D. , Bland , B. , Mann , H. , Brown , V. and Ross , D. ( 1999 ) Personal exposure to air pollutants in Herts, England . Proceedings of the 8th International Conference on Indoor Air Quality and Climate (Indoor Air ’ 99) , Edinburgh, 8 – 13 August 1999 , Vol. 5, pp. 288 – 93 .

Crump , D. , Raw , G. , Upton , S. , Scivyer , C. , Hunter , C. and Hartless , R. ( 2002 ) A protocol for the assessment of indoor air quality in homes and offi ce buildings . BRE Report BR 450 , CRC Ltd , London, UK .

Crump , D. , Brown , V. , Rowley , J. and Squire , R. ( 2004 ) Reducing ingress of organic vapors into homes situated on contaminated land . Environmental Technology , 25 , 443 – 50 .

Crump , D. , Brown , V. , Carson , A. and Harrison , P. ( 2007 ) Assessment of risk from inhalation exposure to benzene – a case study . Proceedings of the Tenth Annual UK Review Meeting on Outdoor and Indoor Air Pollution Research , 1 – 2 May, IEH, Cranfi eld University, 1 – 2 May 2007 .

Dimitroulopoulou , S. , Crump , D. , Coward , S. , Brown , V. , Squire , R. , Mann , H. , White , M. , Pierce , B. and Ross , D. ( 2005 ) Ventilation, air tightness and indoor air quality in homes in England built after 1995 . BRE report BR 477 , BRE , Watford, UK .

Dingle , P. , Tan , R. and Cheong , C. ( 2000 ) Formaldehyde in occupied and unoccupied caravans in Australia . Indoor and Built Environment , 9 , 233 – 6 .

Directive 96/62/EC ( 1996 ) Directive of the European Parliament and the Council of 27 September 1996 on Ambient air quality assessment and management . Offi cial

Journal of the European Communities , L296 , 55 .

ECA ( 1994 ) Sampling strategies for volatile organic compounds (VOCs) . European collaborative action – Indoor Air Quality and its Impact on Man, Report No 14, European Commission, EUR 16051 EN, Luxembourg.

EN ( 1995 ) 838 . Workplace Atmospheres – Diffusive Samplers for the Determination of Gases or Vapors – Requirements and Test Methods , The British Standards Institution , London, UK .

EN ( 2004a ) 13528 - 2 . Diffusive Samplers for the Determination of Concentrations of Gases and Vapors; Specifi c Requirements and Test Methods , The British Standards Institution , London, UK .

EN ( 2004b ) 14412 . Indoor Air Quality – Diffusive Samplers for the Determination of Concentration of Gases and Vapors – Guide for Selection, Use and Maintenance , The British Standards Institution , London, UK .

EN ( 2005a ) 14662 - 5 . Ambient Air Quality – Standard Method for Measurement of Benzene Concentrations; Diffusive Sampling Followed by Solvent Desorption and Gas Chromatography , The British Standards Institution , London, UK .

EN ( 2005b ) 14662 - 4 . Ambient Air Quality – Standard Method for Measurement of Benzene Concentrations; Diffusive Sampling Followed by Thermal Desorption and Gas Chromatography , The British Standards Institution , London, UK .

EN ISO ( 2003 ) 16017 - 2 . Indoor, Ambient and Workplace air – Sampling and Analysis of Volatile Organic Compounds by Sorbent Tube/Thermal Desorption/Capillary Gas Chromatography – Part 2: Diffusive Sampling , The British Standards Institution , London, UK .

Franklin , P. , Dingle , P. and Dick , S. ( 2000 ) Formaldehyde exposure in homes is associated with increased levels of exhaled nitric oxide in healthy children . Proceedings of Healthy Buildings 2000 , August 6 – 10, Espoo, Finland, Vol. 1, pp. 65 – 70 .

Fujii , M. , Shinohara , N. , Lim , A. , Otake , T. , Kumagai , K. and Yanagisawa , Y. ( 2003 ) A study of the emission of phthalate esters from plastic materials using a passive fl ux sampler . Atmospheric Environment , 37 , 5495 – 504 .


Haruki , O. , Yasuhiro , M. , Kazuaki , B. and Hironori , S. ( 2003 ) Field measurement of formaldehyde in government offi ces . Proceedings of Healthy Buildings 2003 , December, 7 – 11, Singapore , pp. 554 – 9 .

Hafkenscheid , T. , Langelaan , F. , Baldan , A. and Wilkinson , G. ( 2001 ) Determination of uptake rates for 4 - week diffusive sampling of chlorinated butadienes in indoor air . Proceedings of International Conference on Measuring Air Pollutants by Diffusive Sampling, Montpellier , France, 26 – 29 September 2001 , pp. 229 – 34 .

Hoffman , K. , Krause , C. , Seifert , B. and Ullrich , D. ( 2000 ) The German environmental survey 1990/92 (GerES11): sources of personal exposure to volatile organic compounds . Journal of Exposure Analysis and Environmental Epidemiology , 10 , 115 – 25 .

HSE ( 1992 ) n - Hexane in Air – Laboratory Method Using Charcoal Diffusive Samplers, Solvent Desorption and Gas Chromatography , MDHS 74 , Health and Safety Executive , Bootle, UK .

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Jia , C. , Batterman , S. , Godwin , C. and Hatzivalis , G. ( 2005 ) Distributions of volatile organic compounds (VOCs) in indoor and outdoor air among industrial, urban and suburban neighbourhoods .

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Kirchner , S. , Gauvin , S. , Golliot , F. , Ramalho , O. and Pennequin , A. ( 2003 ) French permanent survey on indoor air quality – microenvironmental concentrations of volatile organic compounds in 90 French dwellings . Proceedings of Healthy Buildings 2003 , December, 7 – 11, Singapore , pp. 349 – 54 .

Mann , H. , Crump , D. and Brown , V. ( 2001 ) Personal exposure to benzene and the infl uence of attached and integral garages . Journal of the Royal Society for the Promotion of Health , 121 ( 1 ), 38 – 46 .

Modig , L. , Sunnerson , A. , Levin , J. , Sundgren , M. , Hajenbjork - Gustafsson , A. and Forsberg , B. ( 2004 ) Can NO 2 be used to indicate ambient and personal levels of benzene and 1,3 - butadiene in air . Journal of Environmental Monitoring , 6 , 957 – 62 .

Mohle , G. , Crump , D. , Brown , V. , Hunter , C. , Squire , R. , Mann , H. and Raw , G. ( 2003 ) Development and application of a protocol for the measurement of indoor air quality . Indoor and Built Environment , 12 ( 3 ), 139 – 50 .

Ni , Y. , Kumgai , K. , Yoshino , H. and Yanagisawa , Y. ( 2005 ) A pilot study on VOCs and carbonyl compounds in Chinese residences: overall results of 8 cities . Proceedings of Indoor Air 2005 , September, 4 – 9, Beijing, China , pp. 567 – 71 .

Otson , R. , Fellin , P. and Tran , Q. ( 1994 ) VOCs in representative Canadian residences . Atmospheric Environment , 28 ( 22 ), 3563 – 9 .

Raw , G. , Coward , S. , Brown , V. and Crump , D. ( 2004 ) Exposure to air pollutants in English homes . Journal of Exposure Analysis and Environmental Epidemiology , 14 , S85 – S94 .

Rehwagen , M. , Schlink , U. and Herbarth , O. ( 2003 ) Seasonal cycle of VOCs in apartments . Indoor Air , 13 , 283 – 91 .

Stymme , H. , Emenius , G. and Boman , C. ( 2005 ) Long term ventilation variation in two naturally ventilated Stockholm dwellings . Proceedings of Indoor Air 2005 , September, 4 – 9, Beijing, China , pp. 3239 – 43 .

Uchiyamaa , S. , Aoyagi , S. and Ando , M. ( 2004 ) Evaluation of a diffusive sampler for measurement of carbonyl compounds in air . Atmospheric Environment , 38 ( 37 ), 6319 – 26 .

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Volatile organic compounds comparison of personal exposure and indoor air quality measurements . Proceedings of Indoor Air ’ 96 , July 21 – 26, Nagoya, Japan, 1996 , Vol. 4, pp. 301 – 6 .

Van den Hoed , N. and Van Asselen , O. ( 1991 ) A computer model for calculating effective uptake rates of tube - type diffusive

samplers . Annals of Occupational Hygiene , 35 ( 3 ), 273 – 85 .

Venn , A. , Cooper , M. , Antoniak , M. , Laughlin , C. , Britton , J. and Lewis , S. ( 2003 ) Effects of volatile organic compounds, damp and other environmental exposures in the home on wheezing illness in children . Thorax , 58 , 955 – 60 .

Real - Time Monitoring of Indoor Organic Compounds Yinping Zhang and Jinhan Mo

4

4.1 Introduction

In general, tests for indoor air quality are categorized as either fi eld tests or labora-tory tests. Controlling indoor gas - phase chemical contaminants involves three aspects: pollution source control, contaminant dispersion control (including ven-tilation dilution and air fl ow organization) and air cleaning techniques. For both fi eld tests and laboratory tests or any of the aspects of controlling indoor air quality, real - time monitoring of organic compounds is very important, particularly to show temporal indoor contaminants and their continuous concentration variations.

In the fi eld tests for determining indoor air quality or human exposure to harmful pollution, real - time monitoring is necessary because the pollution con-centration tends to vary greatly over time and space (Ezzati, Mbinda and Kammen, 2000 ; Lai et al ., 2004 ; Lirk, Bodrogi and Rieder, 2004 ; Stefaniak et al ., 2000 ).

The available research on indoor VOC source/sink behavior falls into two catego-ries: experimental studies and modeling. The literature concerned with the experi-mental approach to studying the characteristics of VOC emissions from building materials is extensive, and includes descriptions of various methods such as the chamber method (ASTM, 1997, 2001 ), the FLEC method (Wolkoff, 1996 ), the CLIMPAQ method (Gunnarsen, Nielsen and Wolkoff, 1994 ) etc. In most cases a real - time technique of monitoring the instantaneous VOC concentration in the tested space is needed. Even in the modeling research such as Little, Hodgson and Gadgil (1994) , Cox, Little and Hodgson (2001a, 2001b) , Guo (2002a, 2002b) , Xu and Zhang (2003, 2004) , Deng and Kim (2004) etc., the determination of key parameters of VOC emissions such as the diffusion coeffi cient, partition coeffi -cient etc. is based upon measurements some of which also require accurate real - time VOC concentration measurement.

The real - time monitoring of some species concentration distribution in indoor air is an effective way to evaluate the dispersion of fresh air and/or gas - phase chemical pollutants, for example when determining the ventilation effectiveness, the air exchange rate ( ACH ) or the air age distribution for a given space (Park et al ., 1998 ; Shin et al ., 2005 ). It also provides a helpful validation for transient


65

66 4 Real-Time Monitoring of Indoor Organic Compounds

computational fl uid dynamics ( CFD ) techniques (Demokritou et al ., 2002 ; Kobayashi and Chen, 2003 ). Air - cleaning is another technique to guarantee clean indoor air. Accurate evaluation of the air cleaner performance such as cleaning effectiveness, life span etc. for chemical gas pollution is also based upon real - time monitoring techniques (Grinshpun et al ., 2005 ).

The objective of the present article is to introduce both basic and advanced real - time monitoring techniques to researchers or engineers in the fi eld of indoor gas - phase chemical pollutant control. Each real - time monitoring technique is introduced according to: a detection or measurement principle; a measuring method; selectivity, sensitivity and accuracy; application illustrations. Examples of applying these real - time monitoring techniques are also presented.

4.2 Proton Transfer Reaction – Mass Spectrometer ( PTR – MS )

A proton transfer reaction – mass spectrometer ( PTR – MS ) system has been devel-oped which allows on - line measurement of VOCs with concentrations as low as a few pptv ( parts per trillion by volume ) (Hansel et al ., 1998 ). The acute measure-ment sensitivity to VOCs and real - time characteristics make it a very powerful tool in both indoor and outdoor environmental research.

4.2.1 Detection Principles

The detection principle of the PTR – MS is based on reactions of H 3 O + ions, which perform nondissociative proton transfer to most of the common VOCs but do not react with any of the components present in clean air (Knighton and Grimsrud, 2003 ).

The PTR – MS system consists of three sections: an ion source, a drift tube ( DT ) and an ion detection section, as shown in Figure 4.1 . H 3 O + ions (primary ions) are produced at high concentration from pure water vapor within a hollow cathode ( HC ) ion source and pass via a Venturi - type inlet ( VI ) into the DT. The analyzed air sample is introduced into the DT at a fl ow rate of about 11 ml min − 1 under a pressure of about 2 mbar. Because of their low proton affi nities (approximately 180 – 600 kJ mol − 1 ), the major components of air (N 2 , O 2 , Ar, CO 2 ) undergo nonreac-tive collisions with H 3 O + ions and, therefore, act as a buffer gas. These components are removed by a turbo - molecular vacuum pump. However, any collisions of H 3 O + ions with gaseous constituents, M, possessing a proton affi nity greater than that of water will result in a proton transfer reaction:

H O M MH H O3 2+ ++ → + (4.1)

The proton affi nity of some gases found in indoor air is shown in Table 4.1 . From the table it is seen that most gases in ambient air (e.g., O 2 , N 2 , CO 2 ) cannot take

4.2 Proton Transfer Reaction – Mass Spectrometer (PTR–MS) 67

Figure 4.1 Schematic of PTR – MS system. HC : hollow cathode ; SD: source drift region; VI : Venturi - type inlet (Hansel et al ., 1998 ).

Table 4.1 The proton affi nity of some indoor air gases.

Name Molecular formula Proton affi nity (kJ/mol)

Oxygen O 2 421.2 Nitrogen N 2 494.0 Carbon dioxide CO 2 540.9 Carbon monoxide CO 593.3 Water H 2 O 691.7 Formaldehyde CH 2 O 713.4 Benzene C 6 H 6 750.7 Toluene C 7 H 8 784.6 Ethylbenzene C 8 H 10 788.4 p - Xylene C 8 H 10 795.5

part in the proton transfer reaction because they have a lower proton affi nity than water. However, most of the organic compounds in indoor air (e.g., formaldehyde, benzene and toluene) can take part in the proton transfer reaction because of their higher proton affi nity.

The resultant ions (both primary and produced) are mass - selected using a quad-rupole mass analyzer and measured as count rates by an electron multiplier detec-tor. Count rates of the MH + species are subsequently converted to ionic densities and then to mixing ratios of constituent M after consideration of instrumental transmission coeffi cients, temperature, and DT pressure. Instrumental accuracy, which is largely determined by the uncertainties associated with the reported proton transfer reaction rate coeffi cients ( k ), is estimated to be better than 30% (Hayward et al ., 2002 ; Lindinger, Hansel and Jordan, 1998 ).


These exothermic proton - transfer reactions occur on every collision with well known rate constants, having typical values 1.5 × 10 − 9 cm 3 s − 1 < k < 3 × 10 − 9 cm 3 s − 1 . An additional advantage of using primary H 3 O + ions is that many of their proton - transfer processes are nondissociative, so that only one product ion species occurs for each neutral reactant. In order to allow for an accurate quantifi cation of the neutral reactants from measured primary and product ion signals, the reactions of H 3 O + with the neutrals must occur under well - defi ned conditions. This is assured in the PTR – MS system by allowing the H 3 O + reactions to proceed within a DT (Hansel et al ., 1998 ).

Because PTR – MS uses a low - energy ‘ soft ’ ion source, most (although not all) proton transfer processes are nondissociative. As a result, many reactive constitu-ents in the atmosphere produce only one molecule, which is measured as the molecular mass, plus one. The count rates for selected user - defi ned ions (masses) are successively measured using predetermined dwell times for each, enabling longer dwell times to be applied where expected concentrations are low and vice versa. For example, a dwell time of 20 s may be appropriate where an ambient concentration of only a few tens to a few hundred pptv is expected. In this way, experimental precision can be increased for VOCs at low concentrations, although temporal resolution between successive measurements of the same mass may be compromised, particularly when a large number of masses are being analyzed. Additionally, lengthy dwell times may mask any concentration changes that occur on short timescales. In summary, it can therefore be seen that the time taken between successive measurements of the same mass is dictated by the number of masses under consideration and the dwell time of each (Hayward et al ., 2002 ).

4.2.2 Measuring Method

Using a mass fl ow controller, ambient air from a shared inlet line was sampled into the PTR – MS instrument at a fl ow rate of 15 ml min − 1 . The air fl owed through the DT at a pressure of 2 mbar and ambient temperature.

Under conditions where the reactively neutral M, is present at trace levels, the density of the product ion MH + is given by:

MH H O M+ +[ ] = [ ] ⋅[ ]⋅ ⋅3 0k t (4.2)

where [H 3 O + ] 0 and [MH + ] are the densities of the primary reagent ions and the resultant ions respectively; k is the reaction rate constant for the proton transfer reaction (4.1); and t is the reaction time.

Neutral molecule densities were calculated from the measured product ion intensities using the appropriate reaction rate constants and the measured drift times of H 3 O + (1.06 × 10 − 4 s) and H 3 O + (H 2 O) (1.11 × 10 − 4 s) ions with the following equation,


MH O H O H O

[ ] =⋅ ( )⋅ + ⋅ ( )⋅

+

+ + + ++ +

I

I K t I K tMH

H O c H O H O H O c H O H O3 3 3 2 3 23 3 2

(4.3)

where, [M] is the density of reactive neutral molecule; I is the measured product ion intensity; K c is the reaction rate constant for the proton transfer reaction; and t is the reaction time (Knighton and Grimsrud, 2003 ).

Volumetric mixing ratios are determined by dividing the sample densities, from Equation 4.3 , by the drift gas total density and are reported as ppbv ( parts per billion by volume ).

The quadrupole mass spectrometer is scanned at 0.5 s/amu (amu: atomic mass unit) repetitively and continuously over the sampling period. A switching valve is programmed to repeatedly admit ambient air directly into the instrument and then to redirect the sampled ambient air through the catalytic scrubber, which provides a hydrocarbon free gas stream to evaluate the background ion intensities. Active ions are identifi ed by the characteristic intensity pattern created by the repeated sampling of the ambient and scrubbed air samples and the fi nal reported concen-trations are determined as the concentration difference between the ambient and scrubbed air samples (Knighton and Grimsrud, 2003 ).

4.2.3 Accuracy, Linearity, Limits of Detection and Precision

Hayward et al . (2002) studied the performance of the PTR – MS, including accuracy, linearity, limits of detection and precision, experimentally. Figure 4.2 presents the PTR – MS - determined concentrations of isoprene and its mono - substituted 13 C analog following sequential dilutions of a gaseous standard. The standard concen-tration was confi rmed by GC – FID analysis to be 115 ppbv, and the fi nal dilution stage produced a calculated concentration of 99 pptv, an overall dilution of greater than three orders of magnitude. Several observations are noteworthy.

First, the PTR – MS accurately measures the test - gas concentration across almost the entire range of concentrations, with most values lying close to or within 10% of the predicted concentrations. This is comparable to observations made by Hansel et al . (1998) who carried out a similar experiment with benzene and toluene gaseous standards. Furthermore, the value of 10% is close to the lower range of the accuracy of the procedure used to standardize the test gas and is well within the 30% uncertainty range given by Lindinger, Hansel and Jordan (1998) .

Second, it is notable that the data points of the highest 11 concentrations for mass 69 follow a slightly different slope to the remaining data points. This anomaly is most likely due to inaccuracies brought about by the dilution procedure and not a problem with the PTR – MS, since it coincides with a signifi cant change in the confi guration of the mass fl ow controllers used for dilution. It is also worth noting that the anomaly is present in the mass 70 13 C isotopic trace (for the higher con-centrations, at least), providing supporting evidence that the apparent loss of lin-earity is simply due to systematic inaccuracies in the gas - dilution stage.


Third, it is evident that a clear relationship between mass 69 and mass 70 is present. At the higher concentrations, there is very good agreement between the predicted mass 70 concentrations (calculated by applying the natural isotopic ratio to the observed mass 69 concentrations and displayed as an unbroken line) and the observed mass 70 observations. This relationship starts to show notable devia-tions ( > 20%) below isoprene concentrations of 1.2 ppbv, which is equivalent to a mass 70 concentration of just 49 pptv. A similar deviation continues to be observed in concentrations as low as 19 pptv (mass 70), beyond which the deviation becomes considerable. This technique demonstrates the potential of the PTR – MS to measure trace gases at extremely low concentrations with some degree of accuracy. While it is diffi cult to reliably produce test gases at very low concentrations, an interpretation of the ratios of isotopic analogs provides a convenient surrogate.

Finally, the magnitude of the instrumental noise (precision) is observed to be proportional to the square root of the recorded signal (i.e., as the signal increases, so does the associated noise).

Instrumental accuracy is further highlighted by Figure 4.3 , which shows the PTR – MS determination (in both counts - per - second and ppbv) of the output from a dimethyl sulfi de ( DMS ) permeation device with and without an in - line charcoal fi lter. The device was gravimetrically found to produce a concentration of 2.60 ppbv

Figure 4.2 Measured concentrations of isoprene (C 5 H 8 , molecular mass 68) at protonated mass 69 (open circles) and its monosubstituted 13 C analog at protonated mass 70 (closed circles) following a stepwise dilution of a 115 ppbv gaseous isoprene standard with zero - air. Each data point represents a mean of between 20 and 40

values. The dashed lines represent a 1 : 1 ratio between the gravimetric and PTR – MS determinations with 10% offset. The unbroken line represents the predicted 13 C analog concentration based upon the observed molecular mass 69 values, assuming a natural isotopic ratio of 17.7/1 (Hayward et al ., 2002 ).


(5%) at 30 ° C. This mixing ratio compares very favorably with the mean blank - subtracted PTR – MS prediction of 2.79 ppbv, close to the uncertainty range of the gravimetric determination. In addition, the low drift of the instrument noise for this mass (across 4 h) is demonstrated. This feature has positive implications for the accuracy with which measured concentrations can be quoted, particularly at low concentrations. It also demonstrates that high - frequency instrument blank measurements are not necessary with the PTR – MS, enabling a higher temporal resolution of sampling.

Figure 4.3 illustrates that the magnitude of instrument noise increases as the instrument signal increases. Because the instrument noise signal, measured as the number of counts ( v ) per fi xed unit time, occurs at a random but defi nite average rate, it can be considered as being distributed according to the Poisson distribution. For a suffi ciently large mean number of counts ( v ) per fi xed time (e.g., v > 9), the Poisson distribution is well - approximated by the Gaussian distri-bution, with both the mean and variance equal to v . It therefore follows that standard deviation ( SD ) is defi ned as the square root of v , and it can be concluded that the instrument noise is proportional to the square root of the mean signal.

As noted previously in the text and elsewhere (Holzinger et al ., 2001 ), instru-ment precision can be improved with longer dwell times (i.e., the fractional uncer-tainty, defi ned as the SD of the signal divided by the mean signal, decreases as v increases). However, such an ‘ averaging ’ technique may not always be suitable (e.g., when VOC concentrations, i.e., count rates, are high but fl uctuating). There-fore, it was felt necessary to quantify instrument noise so that dwell time could be optimized, and degrees of signifi cance could be attributed to changes in the

Figure 4.3 Measured concentrations and raw counts per second (cps) of a gaseous dimethyl sulfi de (DMS) standard produced using a permeation device and calibration oven (Hayward et al ., 2002 ).


PTR – MS output. To achieve this, the instrumental response (cps) to constant concentrations of DMS (protonated mass 63), isoprene (69), and its 13 C isotopic analog (70) was monitored. A range of instrument dwell times was used (from 0.1 to 20 s), and the gaseous concentrations were varied between an instrument blank (equivalent to 50 pptv prior to blank subtraction) and 41 ppbv. For each given set of gaseous and instrumental parameters, at least 170 data points were analyzed in order to assess the suitability of a noise statistic ( NS ), developed from the Poisson distribution described previously, in modeling instrument noise under all condi-tions. The noise statistic effectively enables the SD associated with any mean signal for any dwell period to be reliably predicted and is defi ned as

NS =( )

( )⋅ ( )mean signal cps

mean signal cps dwell s (4.4)

Hayward et al . (2002) summarized the results obtained and tabulated the percent-age of observed values (cps) that lie outside of the mean signal (cps) (0.5 × NS, 1 × NS, and 2 × NS, analogous to standard deviations). The mean of these values are 31.3%, 15.9%, and 1.98%, respectively, which are very close to the Gaussian approximated predictions of 30.9%, 15.9%, and 2.2%. This striking similarity illustrates that the instrument noise is well - described by the Gaussian approxima-tion when v is suffi ciently large and provides further evidence that instrument noise is not biased in any way. As would be expected, the percentage of values lying outside of the given range decreases as the NS range increases. These values appear to be independent of the protonated mass, dwell time (s), and count rate (cps), indicating the suitability of the noise statistic in modeling the noise compo-nent of any instrumental signal, regardless of the aforementioned factors. This enables us to readily predict what change in instrumental count rate (cps) would be required in order for us to be certain that the change was not just a factor of noise but a genuine change in concentration. For example, referring to the DMS data given in Figure 4.3 (mean: 102.9 cps, dwell: 10 s), we can predict that approxi-mately 96% of the noise signal will lie within 102.9 (2 × NS) cps (i.e., on average, 2% of the observed values will lie above and 2% will lie below this range), which in this case equals 96.5 – 109.3 cps (3.2 – 3.6 ppbv prior to mean blank subtraction). Therefore, a data point lying above or below this range would have, at most, a 2% chance of occurring simply as a result of noise. It is referred as the ‘ 2% level ’ . The further this point lies above or below the mean (2 × NS), the greater the probability that it is not a factor of noise, but representative of a genuine change in atmo-spheric concentration.

4.2.4 Applications of PTR – MS

Instrumental monitoring of the gaseous components of indoor air or indoor air samples has become routine in environmental analysis. In all monitoring systems, a number of characteristics are considered essential. Among these are instrumen-

4.3 Photo-acoustic Spectroscopy 73

tal accuracy and precision, a high temporal sampling resolution coupled with a low detection limit and a high signal - to - noise ratio (sensitivity). Most instrumental monitoring procedures possess several of these attributes, yet few are able to offer all, often compromising one attribute in favor of another (e.g., high temporal reso-lution is often sacrifi ced during a pre - concentration stage in order to increase the sensitivity of a sampling system).

The PTR – MS claims to possess all of the aforementioned attributes for real - time monitoring of VOCs, combining a high degree of sensitivity with relatively short integration periods (as compared to existing sampling and analytical systems) for a suite of volatile organic species found at trace concentrations in air.

The PTR – MS has been developed and used for a number of different applica-tions. For example, Lindinger, Hansel and Jordan (1998) described the monitoring of metabolic processes in the human body by analyzing breath; the monitoring of fruit and meat aging; and the decay of vegetation. This monitoring is done by PTR – MS analysis of VOCs emitted, and the analysis of trace components of ambient air. Crutzen et al . (2000) , Warneke et al . (2001) , Williams et al . (2001) , and Poschl et al . (2001) described measurements made by PTR – MS from an aircraft over the tropical rainforests of Surinam.

Hayward et al . (2002) demonstrated that PTR - MS could reliably measure a wide range of VOCs and with a time resolution suffi ciently fast to capture the dynamics of many environmental processes (e.g., the light dependency of isoprene emis-sions from vegetation). They also demonstrated that the components of the instru-ment output (signal plus noise) were easily characterized, enabling a simple interpretation of measurements.

4.3 Photo - acoustic Spectroscopy

Photo - acoustic spectroscopy ( PAS ) is a kind of infrared ( IR ) spectroscopy which is a popular choice for real - time monitoring of VOCs at ppbv levels. Recently there has been a great revival of interest in PAS because it offers much greater sensitivity than conventional spectroscopic techniques. All spectroscopic methods yield quan-titative and qualitative information by measuring the amount of light a substance absorbs; PAS simply measures this in a more sensitive way.


Many gas - phase chemical pollutants such as VOCs absorb IR light at specifi c wavelengths. When a broad spectrum of IR light is passed through a gas, some of the frequencies are absorbed while the rest are transmitted without being absorbed. Those frequencies absorbed correspond to the natural frequencies of the vibration modes of the gas molecules or to a harmonic of these vibrations. The amount of energy the gas molecules absorb is proportional to its concentration


and it is immediately released as heat which causes the pressure to rise. When the incident light is modulated at a given frequency, the pressure increase is peri-odic at the modulation frequency. The pressure waves or sound waves are easily measured with a microphone. The intensity of the sound emitted depends on the nature and concentration of the substance and the intensity of the incident light. This phenomenon is known as the photo - acoustic effect.

A condenser microphone consists of a thin metallic membrane in close proxim-ity to a rigid backplate. This forms an air dielectric capacitor, the capacitance of which varies with the distance between the plates. A small equalization tube vents the inner cavity of the microphone to the static atmospheric pressure in the mea-surement chamber. The vent is large enough to permit equalization of the static atmospheric pressure without affecting the sound pressure variations within the measurement chamber. When the pressure in the measurement chamber increases or decreases, the fl exible membrane moves in or out synchronously. It results in the capacitance between the plates varying accordingly. The capacitance is measured by applying a fi xed charge to the microphone and measuring the associated voltage change. The alternating current (AC) voltage is an exact replica of the sound pressure variations in the measuring chamber.

Each substance has a unique IR spectrum. The spectra of almost all known compounds are compiled in various reference volumes and the identity of a sub-stance can be confi rmed by comparison of its spectrum with these standard spectra. Well - resolved spectra yield detailed structural information such as bond lengths.

The strength of the absorption is proportional to the amount of absorbing species present. By calibrating with a standard sample of known concentration, the concentration can be determined.

4.3.2 Measuring System and Method

A photo - acoustic system includes the following essential components: a chamber to contain the gas sample, a light source, some means of modulating the light (usually a chopper), a detector to measure the sound (usually a microphone), and some methods of processing the signal (the level of sophistication of the signal analysis depends entirely on the requirements), shown in Figure 4.4 .

4.3.2.1 Discrete Sampling: Nondispersive PAS The nondispersive (fi lter - based) PAS detector consists of very similar components to the original setup used by Alexander Bell: an IR light source, a chopper wheel and a measurement cell. In addition, optical fi lters have been added to improve selectivity, as has a pump to introduce the sample into the measure-ment cell.

The IR light source is placed in a parabolic mirror in order to concentrate and focus the IR light inside the measurement cell. The chopper wheel, rotating with a well - defi ned frequency, will modulate the light, generating light - pulses which


enter the measurement cell. The optical fi lter will ensure that only light at the selected wave - length enters the measurement cell.

The instrument works in a semi - continuous way. First, the pump purges the sample lines and the measurement cell in order to fl ush out the old sample and bring in the new one. The valves to and from the measurement cell are then closed and measurement starts: the IR source is turned on, the chopper wheel starts rotating and the microphones pick up the photo - acoustic signal. The optical fi lter wheel positions each of the optical fi lters in the light path, one after the other, until all fi lters have been measured. Finally the instrument calculates the concen-tration of each gas, the results are displayed and the whole procedure starts all over again.

One optical fi lter for the measurement of water vapor is installed in the fi lter wheel as the standard. In this way, the water vapor is measured in each sample, enabling the instrument to subtract the absorption signal coming from water vapor from the measured signal for a gas of interest, giving a more correct mea-surement. Similarly, by measuring the same air sample with more optical fi lters, one fi lter selected for each gas in the sample, it is possible to compensate for the interference of one gas on another.

Figure 4.4 A typical set - up of photo - acoustic system (INNOVA Air Tech., 1997 ).


The entire measurement system is very compact (the volume of the measure-ment cell is 3 cm 3 ) with a very small internal volume in the tubing and pump. Besides making the instrument very compact, it allows accurate measurements to be performed on even very small sample volumes. This, together with the out-standing stability of the PAS detector and the ability to compensate for interfer-ence, is the biggest advantage of the system.

4.3.2.2 Discrete Sampling: FTIR/PAS For analyzing unknown gas samples in order to identify the components and to quantify them, the Fourier Transform IR analyzer ( FTIR ) has proven to be a pow-erful tool, particularly in laboratories. Combining the FTIR technique with the PAS detector, the small but very rugged and stable detector opens up new dimen-sions in FTIR by making it possible to build a portable FTIR analyzer.

Basically an FTIR analyzer is based on the same principles as a nondispersive IR monitor. But the chopper wheel and the optical fi lters have been substituted by a so - called Michelson interferometer. The interferometer consists of some mirrors and a beam - splitter. By splitting the IR light into two beams: one which is refl ected by a fi xed mirror and one which is refl ected by a moving mirror, and then combining the two beams again in the measurement cell, the combined beam is now modulated with a different modulation frequency for each wave-length of the light. In this way the entire IR range is ‘ scanned ’ and the absorption at each wavelength is then measured in the PAS detector. In other words it cor-responds to measuring the gas sample with an infi nite number of optical fi lters. Therefore, the PAS not only identifi es the various components of the sample but also measures the concentration of each component.

4.3.2.3 Continuous Flow - PAS The PAS detector can also be used in a continuous fl ow system. By eliminating the inlet and outlet valves to the measurement cell, and pumping the sample through the cell at a constant, well - controlled rate of fl ow, the continuous fl ow PAS system can perform real - time monitoring.

In this system the optical fi lters are combined to form the optical window through which the IR light enters the cell. Passing the chopper wheel modulates the light, but this chopper wheel is perforated at different distances from the center. When rotating the chopper wheel at constant velocity, three different modulation frequencies are obtained – corresponding to the three optical fi lters. A microphone picks up the photo - acoustic signal downstream.

Due to the constant fl ow through the measurement cell the response time is extremely short. However, the constant fl ow creates turbulence in the cell, which the microphone detects as noise. Therefore the detection limit in a continuous fl ow system is higher than in a discrete sampling system.

As a special feature the PAS detector can be combined with a magneto - acoustic detector ( MA ) for the measurement of oxygen (O 2 ). Oxygen does not absorb IR light but it is paramagnetic. This means that if a switched magnetic fi eld is applied


to the gas stream, oxygen will vibrate and the microphone will pick up the vibra-tions. By comparing the oxygen measurement with a reference gas (normally ambient air where O 2 is constantly 20.95% vol.), the MA detector can measure with an accuracy which can normally only be achieved with far more expensive Mass Spectrometers. As an example, this is particularly useful in Fermentation Monitoring where metabolic oxygen consumption is an important indicator of the microbial activity.

4.3.3 Selectivity, Sensitivity and Accuracy

As mentioned earlier, PAS is an accurate technique as it measures absorption directly. The PAS method has additional advantages over conventional transmis-sion spectroscopy. It uses a very stable transducer (microphone) so calibration is seldom required more than 4 times a year (as opposed to daily/weekly).The linear response of the microphone enables measurement of gas concentrations over a wide dynamic range without having to change the range setting or to recalibrate. It uses a small volume of gas (about 3 cm 3 ) in the measurement cell (about 3 – 4000 cm 3 ) – thus reducing the time between measurements (i.e., the response time is fast).

Due to the very linear response of a precision microphone, the response of a photo - acoustic cell is linear over a wide dynamic range, typically four to fi ve orders of magnitude (e.g., from 1 ppm to 10 4 or 10 5 ppm). This allows the same instru-ment to be used for monitoring both the trace amounts of pollutants, which are generally found in ambient air, and high concentrations, which are found at the source.

It is essential that toxic gas concentrations are measured accurately. The low range - drift of a PAS monitor enables accurate measurements to be performed over long periods of time. Typical reproducibility is 1% of the measured value with a range - drift of less than 2.5% of the measured value in three months.

In contrast to many other monitoring systems, the measuring chamber of a photo - acoustic monitor is sealed. Hence, the air must be analyzed as discrete samples rather than continuously. Nonetheless, this does not adversely affect the response time because small cell volumes, typically 3 cm 3 can be used due to the high sensitivity available. Small discrete samples can be taken frequently and the true concentration of the ambient air measured.

Another aspect which affects the response of a detector is the ease with which it can be purged. Many nonspectroscopic instruments have a long recovery time after a high concentration is detected – sometimes as much as one hour. The time taken to purge the cell and take a new measurement is typically less than 1 minute for a PAS based instrument.

According to the detection principle of PAS, the interference between two gases sometimes is unavoidable. A monitor, which is able to detect several gases in a single sample, can reduce, and sometimes overcome, the problem by cross -


compensating for the presence of the interfering species. From these measure-ments it is possible to calculate the gas B ’ s infl uence on gas A ’ s measurement and thus fi nd gas A ’ s true concentration.

Toxic and polluting gases are normally found as trace components in atmo-spheric air. The various gases which are found in normal atmospheric air are therefore potential interferents when detecting toxic and polluting gases. Fortu-nately, however, the main constituents of air do not absorb IR radiation at all, as is illustrated by the simple fact that we get heat from the sun. However, the two minor constituents in air, water and carbon dioxide, do absorb IR light.

Immunity to interfering species is perhaps the most important consideration in any gas detection method. Because gases absorb IR light over a narrow spectral range and only at characteristic wavelengths, instruments based on IR absorption spectroscopy offer greater selectivity than most other techniques.

It sometimes happens that the most likely interference is of a very similar struc-ture to the gas of interest. Consequently, their spectra are also very similar and there is considerable overlap of absorption bands. In such cases, high selectivity can still be obtained by irradiating at two wavelengths at which the gases have different ratios of absorption.

This is how to cope with any interference caused by the presence of atmospheric water vapor and carbon dioxide. Air samples are irradiated by light at two different wavelengths, one, which corresponds to a strong absorption band of water vapor and the other of carbon dioxide. From these measurements it is possible to com-pensate each measurement for interference from these two gases. It is of major importance that the water vapor interference compensation is accurate, which is ensured by a thorough calibration procedure.

4.3.4 Applications of PAS

Photo - acoustic Field Gas Monitor 1312 is a gas monitor based on PAS which is accurate, reliable and quantitatively stable. It can be used for on - line concentration monitoring of up to 5 components plus water vapor in gas mixtures. It can also be used to measure the ventilation rate by using tracer - gas and so on (INNOVA Air Tech., 1997 ).

The Gas Monitor 1312 can be used separately, as well as with a multi - channel sampler 1303. Samples of up to 6 locations can be delivered to Photo - acoustic Gas Monitor 1412 and monitored in real - time.

4.4 Flame Ionization Detection

A fl ame ionization detector ( FID ) is a measurement device which is specifi c to hydrocarbons, with a measurement principle based on the combustion of hydro-carbons in a hydrogen fl ame.

4.4 Flame Ionization Detection 79

4.4.1 Detection Principle

In the burning chamber of an FID, a fl ame of highly pure hydrogen in hydrocarbon - free air is burned under controlled conditions (fl ow rates, pressure and temperature) in an electrostatic fi eld, which has a typical gradient of some hundred volts over the burning chamber in parallel to the fl ow. In the inner part of the reducing hydrogen fl ame (pyrolysis zone), the C − C bonds are reduced to C − H. In the outer part of the fl ame (oxidation zone) where suffi cient oxygen is present, C − H is oxidized according to

CH O CHO+ → ++ −e (4.5)

forming an intermediate cation and an electron. The nozzle, injecting the mixture of hydrogen and the sample gas into the burning chamber, is kept at negative potential, and so the electrons are accelerated to a ring - shaped collector anode and produce the signal current. Due to the measurement principle, the signal current is largely proportional to the number of carbon atoms of the measured hydrocar-bon. The signal is also dependent on the species of hydrocarbon because hetero substituted carbon atoms contribute less or not at all to the signal current. This circumstance is represented by a response factor, which is also dependent on the FID, as it is infl uenced for example by the burner geometry and the operation mode (Nikos, 2004 ).

A fast FID consists of a main control unit and two remote sampling heads (which house the FIDs). The dual channel nature of the instrument enables simul-taneous real - time measurement in two locations allowing, for example, evaluation of catalyst performance.

4.4.2 Measuring System and Method

Air and fuel gas (hydrogen) are introduced into the FID to generate a hydrogen fl ame. The sample gas is introduced into the hydrogen fl ame, shown in Figure 4.5 . Any hydrocarbons in the sample will produce ions when they are burnt. Ions are detected using a metal collector which is biased with a high DC voltage. The current across this collector is thus proportional to the rate of ionization which in turn depends upon the concentration of HC in the sample gas.

The ionization process is very rapid, so the slow response time of conventional FIDs is mainly due to sample handling. A typical slow analyzer might have a response time of 1 – 2 seconds. A fast response FID analyzer uses conventional detection principles and a unique patented sampling system to give millisecond response times.


4.4.3 Selectivity and Sensitivity

It is clear that H 2 O or other molecules which cannot be further oxidized by a fl ame do not change the (ionic) current and, therefore, do not show a signal in an FID. Other gases like CO or NO, that can undergo oxidation do not contribute to the signal current due to a different reaction path (Nikos, 2004 ). This makes the FID a reasonable reference method for monitoring VOCs, as it reacts very sensitively toward the expected target analytes and not toward the expected main interfering compound in the air.

FID has no selectivity toward VOCs, in other words, it can ’ t tell VOC ‘ A ’ from VOC ‘ B ’ but a TVOC concentration. The sensitivity of FID is about 0.1 ppmv. However, FID has a good linearity with the response and sample concentration.

4.4.4 Applications of FID

FID is widely used as a detector of GC for organic compounds, but not so popular for real - time use.

Real - time FID may be used in safety control. However, its usage is often limited due to its large size with a hydrogen generator or a hydrogen bottle.

4.5 Photo - ionization Detection

A photo - ionization detector ( PID ) measures VOCs in concentrations from ppbv up to 10 000 ppmv. A PID is a very sensitive broad - spectrum monitor, like a low - level lower fl ammable limit ( LEL ) monitor.

Figure 4.5 Schematic of a fast response FID.

4.5 Photo-ionization Detection 81


A PID uses an ultraviolet ( UV ) light source to break down chemicals to positive and negative ions that can easily be counted with a detector. Ionization occurs when a molecule absorbs the high energy UV light, which excites the molecule and results in the temporary loss of a negatively charged electron and the forma-tion of positively charged ion. The gas becomes electrically charged. In the detector these charged particles produce a current that is then amplifi ed and displayed on the meter as ‘ ppmv ’ or even in ‘ ppbv ’ . The ions quickly recombine past the elec-trodes in the detector to ‘ reform ’ their original molecule (Figure 4.6 ). PIDs are nondestructive; they do not ‘ burn ’ or permanently alter the sample gas, which allows them to be used for sample gathering (RAE Systems, 2005 ).

All elements and chemicals can be ionized, differing only in the amount of energy they require for ionization. The energy required to displace an electron and ‘ ionize ’ a compound is called its ionization potential ( IP ), measured in electron volt s ( eV ). The light energy emitted by a UV lamp is also measured in eV. If the IP of the sample gas is less than the eV output of the lamp, the sample gas will be ionized.

Although benzene, trichloroethylene and other major harmful VOCs are ionized, such major constituents in air as nitrogen, oxygen, carbon dioxide, methane and water have high ionization energy and therefore are not detected (Figure 4.7 ). This is why the PID serves as a VOC detector.


A PID can accurately measure gases and vapors at low ppmv or even ppb levels. However, the PID is not a selective monitor. It has very little ability to differentiate between chemicals. The PID can tell us how much of a gas or vapor is present, but we have to deduce the exact gas or vapor present. When approaching an unknown chemical release, the PID is set to its calibration gas of isobutylene. Once

Figure 4.6 The principle of a photo - ionization detector.


the chemical is identifi ed by means of placard, manifest, waybill or other means, the PID sensitivity can be adjusted to that chemical so that it reads in an accurate scale. For example, if we calibrate on isobutylene and happen to measure a toluene leak of 1 ppmv the PID will display 2 ppmv because it is twice as sensitive to toluene as it is to isobutylene. Once we have identifi ed the leak as toluene, then the PID scale can be set to a toluene Correction Factor and the PID will accurately read 1 ppmv if exposed to 1 ppmv of toluene.

Correction factor s ( CF , also known as response factors) are a powerful tool in the use of PIDs. They are a measure of PID sensitivity to a particular gas. CFs permit the calibration on one gas while directly reading the concentration of another, elimi-nating the need for multiple calibration gases. PID manufacturers determine CFs by measuring a PID ’ s response to a known concentration of target gas.

Correction factors are scaling factors used to adjust the sensitivity of the PID to directly measure a particular gas compared to the calibration gas. The lower the CF, the more sensitive the PID is to a gas or vapor.

Isobutylene has been used to calibrate PIDs because its responsiveness is in the mid - range of PIDs sensitivity. It is relatively easy to obtain and is nontoxic and nonfl ammable at the low concentrations used for calibration. For years PIDs were calibrated with benzene, but because of its carcinogenic properties benzene cali-brations have been phased out. While PIDs are typically calibrated with isobutyl-ene, they can be calibrated with any ionizable gas. For example, if a PID is to be used to measure only vinyl chloride, the PID can be calibrated directly with a know concentration of vinyl chloride.

4.5.3 Applications of PID

A PID can be used in safety control and source detection. It also can be used in environmental soil contamination monitoring, environmental remediation and

Figure 4.7 Some ionization potentials for common chemicals (RAE Systems, 2005 ).

4.6 Metal Oxide Sensors 83

cleanup, determining the level of toxic VOCs in drinking water, soil and water headspace screening, and toxic hazardous waste monitoring.

4.6 Metal Oxide Sensors

Hydrocarbons are well known as good target molecules for semi - conducting metal oxide ( MOX ) based gas sensors. In fact, the use of metal oxides as gas sensors was fi rst developed and introduced for the monitoring of combustible hydrocarbon gases and for a long time this was the primary application where MOX sensors were used. More recently they have also been used for online monitoring in other applications, taking advantage of the ability to change the properties of hydrocar-bons by the use of different metal oxides, various dopants and sophisticated processing or operation techniques. They are selective toward classes of analytes, which generally is a disadvantage; in the case, where one can expect different alkanes it may contribute to the robustness of the correlation between oil content and sensor response (Nikos, 2004 ).

4.6.1 Measuring Principle

Hydrocarbons are the oldest analytes of interest for the mass application of gas sensors based on metal oxide semiconductors, and SnO 2 is the material which has been in use for this purpose for the longest time. This is why it represents the prototype material for numerous investigations of the mechanisms that cause the change of conductivity of the metal oxide.

The state of the SnO 2 surface is dependent on the ambient gas; in the given application it has about 20% O 2 and contains H 2 O and/or oil vapor in varying quantities. Ambient oxygen is adsorbed on the SnO 2 surface; at about 300 ° C the dominant species are ionosorbed O − and O 2 − . The negative charge is provided by free charge carriers, electrons, originating from the conduction band of SnO 2 , so the counter charge is delocalized in the depletion layer, that is, the surface layer of the SnO 2 grain, which is infl uenced by alterations of the surface states. The ionosorption leads to a surface band bending which works against further iono-sorption and results in an equilibrium state with an O − coverage in the range of 10 − 3 – 10 − 5 of a monolayer (Nikos, 2004 ).

The band bending is macroscopically observable in a change of resistance; in case of the n - type semiconductor SnO 2 the resistance increases in comparison to the situation without ambient oxygen.

The interaction with water leads to a decrease of the resistance according to an opposite but comparable charge transfer mechanism, although the reaction mech-anism is not yet fully understood. (Heiland and Kohl, 1988 ) propose two different mechanisms of H 2 O adsorption:

(1) A homolytic dissociation of H 2 O on the surface and a reaction with lattice Sn and lattice O according to


H O Sn O HO Sn O H2 + + → −( ) + ( ) ++ −lat lat lat lat e (4.6)

In this case the hydrogen and the lattice oxygen form a ‘ rooted ’ hydroxyl group which acts as an electron donor to the conduction band. The following equation takes the charge of the lattice oxygen into account

H O O H+ → ( ) +− − −lat lat e2 (4.7)

In contrast to the rooted hydroxyl group, the hydroxyl group coupled to the Sn is called the isolated hydroxyl group and it is assumed that rather than contributing to the conduction process, it changes the oxidation state of the Sn ion.

(2) The second mechanism proposes a heterolytic dissociation and a proton transfer to lattice oxygen. This rooted hydroxyl group changes into an isolated hydroxyl group by forming a bond to a Sn nearby and an oxygen vacancy with a formal charge of + 2. This model contributes two electrons to the conduction band according to

H O Sn O OH Sn2 02 2 2+ + → −( ) + +− −lat lat lat V e (4.8)

Both mechanisms explain the decrease of the resistance with the formation of a rooted or an isolated hydroxyl group out of an O 2 − of the lattice. In both cases it is assumed that the bonding to the Sn does not contribute to the concentration of free charge carriers, which implies that not all the surface tin atoms are in oxida-tion state +4 because otherwise the formation of the Sn − OH bond would need an electron from the conduction band. This assumption is reasonable because tin has two stable oxidation states, +2 and +4, and the most stable surface of tin dioxide, (110), can easily be conditioned to show atoms with both oxidation states. Further-more it is known that defects like vacancies are an essential factor for the perfor-mance of SnO 2 gas sensors and it probably is not realistic to base a mechanism on the situation on a perfect surface. Emiroglu et al . (2001) and Harbeck et al . (2003) proved the formation of rooted and isolated hydroxyl group on the SnO 2 surface in the presence of water, so the fi nal result is clear even if the exact mecha-nism still allows for speculation.

The well - known cross sensitivity between water and hydrocarbons (e.g., lower sensor signal toward methane with increasing humidity (Weimar, 2002 )) implies a competition for adsorption sites respectively, a comparable starting mechanism, and the abstraction of a hydrogen/proton is also the most logical starting point for the heterogeneously catalyzed combustion of hydrocarbons. It is assumed that, analogously to the water dissociation, the carbon fragment is adsorbed at fi rst at a tin atom and then transferred to an oxygen atom. It is not known which oxygen species plays an essential role in this step, even if it is known that ionosorbed oxygen is more reactive. Figure 4.8 shows the steps of the oxidation of hydrocar-bons (exemplifi ed by propane) which are known. Heiland and Kohl (1988) found an intermediate propoxy - like species and a subsequent propanate like species by means of reactive sputtering in vacuum and mass spectrometric detection.

4.6 Metal Oxide Sensors 85

The propanate - like species proves that not all oxygens for the combustion can be ionosorbed ones because two neighboring ionosorbed oxygens are very improb-able due to the Weisz limitation. It has to be stressed, that even if the propane in Figure 4.8 is bonded to lattice oxygen, this is only done in order to simplify the diagram; it is not known which oxygen species plays which role.

Combustion measurements with excess oxygen (carrier gas: synthetic air; 1 bar) showed that under these conditions H 2 O and CO 2 are the only oxidation products detectable using mass spectrometry and IR (Schmid, Barsan and Weimar, 2003 ). This proves that the complex is bonded to the surface throughout the whole reac-tion channel and under these conditions reaction path 2a in Figure 4.8 is the only relevant one. The complete oxidation of the hydrocarbon is not time limited and proves that the oxygen originates from the ambient gas and in an equilibrium condition the oxygen is continuously re - supplied to the lattice, and subsequently to the surface. Figure 4.8 takes this into account, reaction path 2a and 2b include the oxygen balance (+/ − n/2 O 2 ) and end up on a SnO 2 surface without oxygen vacancies. Reaction paths 1 and 2b cannot be observed under atmospheric

Figure 4.8 Mechanism of the reaction of propane on the heated SnO 2 surface as far as known, focusing on the oxidation of carbon. For the sake of simplifi cation, all oxygen is shown as lattice oxygen (Nikos, 2004 ).


conditions. The (rooted) hydroxyl groups, shown at the end of all three paths are in equilibrium with hydroxyl groups due to water adsorption.

Due to the proposed elementary steps of this sensing principle and the signal transduction from the surface reaction to a change of band structure and conse-quentially, to a change of resistance, as described in detail in (Weimar, 2002 ), the dependency of the sensor resistance from the concentration of analyte is logarith-mic and the resistance change is dependent on the baseline value. In order to have a more transferable parameter for the characterization of sensor performance, the sensor signal S is defi ned as:

SR

REX

= 0 (4.9)

where R 0 is the baseline resistance, and R EX is the resistance at exposure to the analyte.


When working with sensors, one of the most important issues is cross - sensitivity. Due to the sensing principle, this notably affects metal oxide gas sensors, espe-cially in the case of measurements performed in real life conditions. To prove real life feasibility, it is necessary to keep as close as possible to the real life conditions of the application. In the present case, the real life conditions are mainly repre-sented by the use of ambient air as a carrier gas, but also by the chosen experi-mental set up.

The response of an MOX sensor is always related to the gas composition as a whole, which may be rather complex. Of course, the sensitivity can be optimized toward a target gas or classes of them, but the sensitivity toward possible interfer-ing analytes cannot be completely eliminated. Furthermore, it is known that, for example, humidity has a considerable effect, not only as an analyte, causing a sensor response but also as the precursor of surface species infl uencing the sensor response toward the analyte. This means that it is desirable to monitor not only the target compound(s) with a reference device, but also all other relevant com-pounds and parameters, which could infl uence the measurement, in order to prove that the sensor response is caused by the parameter of interest, and to reveal measurement artifacts.

Assuming a dependency on interfering parameters, it is important to investigate the extent of their infl uence in order to prove the overall feasibility of the concept, because cross sensitivities and artifacts can cover or spoil the response toward increased oil content. This can certainly happen in the short - term, but also on long - term - basis, for example because ambient air underlies both daily changes as well as seasonal ones. The sensor response can also be infl uenced by the aging of the oil in the compressor, which also represents a real life effect.

4.8 Examples of Investigations Using Real-Time Monitoring 87

Therefore, besides performing reference measurements in parallel to the sensor measurements, an empirical approach should also be followed: the experimental set up should be operated and monitored for a long enough period to provide real life conditions.

The calibration is obtained by measurements of the heater resistances of the sensors at different temperatures, when heated in an oven. The resistance when actively heated is calculated from simultaneous measurements of voltage and current of the sensor heater. In order to obtain more exact results, the resistance of the measurement holders including wires is subtracted from the results for the sensor resistances.

4.7 Air Sampling and Data Recording

Air sampling is often used in real - time indoor air quality monitoring techniques. There are some conditions that should be taken into account for air sampling. One is to never use Tygon sample tubing, because Tygon sample tubing quickly absorbs many chemical vapors. Tefl on tubing will not absorb chemicals but it can get coated. If the Tefl on tubing gets coated with chemicals, it should be cleaned with anhydrous methanol. Flying objects and dust particles must be removed from the air samples by fi lters at the air inlet.

A multi - channel sample instrument or an automatic switching valve can be used when multiple locations or samples need to be monitored simultaneously. For example, such systems have been developed and manufactured which can monitor multiple points from a central location, including both continuous and discrete sampling systems. The systems can be connected to either PC ’ s or process com-puters for further data processing or for activation of process controllers or opera-tion alarms.

4.8 Examples of Investigations Using Real - Time Monitoring

As mentioned in the introduction, real - time monitoring techniques for indoor chemical gas pollutants have many important applications in the fi eld of indoor air quality. To illustrate, several examples of applying real - time monitoring tech-niques to both fi eld measurement and laboratory investigations are presented in this section.

4.8.1 Laboratory Investigations of VOC Emissions from Building Materials

The following is an example of applying an INNOVA gas analyzer, a real - time monitoring system, to study the infl uence of temperature on VOC emissions, by using the so - called C - history method (Zhang et al ., 2007a ).


4.8.1.1 Experimental Principle For a board of building material placed in an air - tight chamber with inner volume V, it is assumed that: the material is homogenous; the initial VOC concentration is uniform; VOC mass transfer is one dimensional; for a given temperature, the partition coeffi cient K, and the diffusion coeffi cient D are constant; the VOC con-centration in the chamber is uniform. By applying the analytical solution for the problem (Xu and Zhang, 2003 ) and the VOC concentration variation in the sealed test chamber continuously measured by INNOVA 1312, the diffusion coeffi cient D and partition coeffi cient can be obtained by fi tting the analytical solution to the experimental data (Zhang et al ., 2007a ).

4.8.1.2 Experimental System The experimental system is shown in Figure 4.9 . The tested samples are placed in an airtight chamber whose volume is 30 l. By using a water bath, the chamber and air temperature can be maintained at the desired temperature. Tests were conducted at four air temperatures: 18 ± 0.5 ° C, 30 ± 0.8 ° C, 40 ± 0.8 ° C and 50 ± 0.6 ° C, with air humidity uncontrolled but in the range 60 ± 8%. After the system reaches thermal equilibrium, a dose of saturated formaldehyde vapor is injected into the chamber and thereafter the instantaneous concentrations of formaldehyde in the chamber are continuously recorded by an INNOVA - 1312 until the equilibrium concentration C ( t equ ) is reached at the equilibrium time, t equ .

There are several reasons for applying a real - time photo - acoustic monitor to measure the instantaneous chamber compound concentrations for this research. First, its sampling volume is small and the air can be returned into the chamber

Figure 4.9 Schematic of the experimental system (Zhang et al ., 2007a ).


after the analysis with no mass loss of air samples, therefore, the air sampling hardly affects the emission of building materials in the chamber; second, the sampling time interval is quite short, about 1 minute or even shorter, which lowers the relative error of determining D and K values because it can record lots of concentration data for a given time period; third, the gas concentration in the chamber varying with time can be monitored and recorded on a computer automatically.

With the real - time monitoring results (Figure 4.10 ), the partition coeffi cients of formaldehyde for four materials at four temperatures can be calculated (Table 4.2 ).

The results of the study show that temperature has signifi cant effect on both the partition coeffi cient and the diffusion coeffi cient of formaldehyde emissions from the four materials tested. For all four materials, the partition coeffi cient decreases while the diffusion coeffi cient increases with increasing temperature.

Figure 4.10 Chamber concentration/time profi les for four kinds of dry building materials: At the beginning of the test, a dose of formaldehyde was injected into the chamber air, the measurement stopped when the chamber formaldehyde concentration hardly changed (Zhang et al ., 2007a ).


Table 4.2 Partition coeffi cient values of the four dry building materials.

Temperature Particle board Vinyl fl oor Medium density board High density board

18 ° C 2.61 × 10 4 9.62 × 10 3 2.10 × 10 4 3.13 × 10 4 30 ° C 1.60 × 10 4 4.84 × 10 3 1.00 × 10 4 2.56 × 10 4 40 ° C 6.69 × 10 3 3.19 × 10 3 2.82 × 10 4 1.32 × 10 4 50 ° C 4.65 × 10 3 2.29 × 10 3 2.17 × 10 4 7.96 × 10 3

4.8.2 Organic Compounds in Outdoor Air

Indoor air quality is infl uenced by outdoor air quality. Therefore, it is helpful to know the chemical gas pollutants in outdoor air for indoor air quality control.

Real - time monitoring is often used to investigate outdoor air as it can record the concentration automatically over a long period. For example, Hayward et al . (2002) carried out some measurements of urban benzene using PTR – MS. Figure 4.11 presents measured results of benzene and NO concentrations made at the Lancaster City Council urban air - pollution monitoring station during 2001. Data were collected over a 7 day period, during which the mean benzene and NO con-centrations were determined to be 790 pptv and 25.5 ppbv, respectively. Benzene was measured using a dwell time of 20 s and at a temporal resolution of about 4 min. Figure 4.11 a shows that a similar diurnal cycle is clearly distinguishable for both benzene and NO for most days; concentrations rise from a nighttime minimum (approximately 1:00 a.m. to 5:00 a.m.) to peak at a late morning maximum (8:00 a.m. to 10:00 a.m.), then briefl y drops to an early afternoon minimum before rising to a late afternoon maximum (4:00 p.m. to 6:00 p.m.). Concentrations then steadily drop to the nighttime minimum. The late morning and late afternoon maxima are further emphasized in Figure 4.10 b, which focuses on data for 1 day only, June 20, 2001. The diurnal cycle clearly refl ects traffi c fl ows around Lancaster city center, peaking during morning and afternoon rush - hour periods, confi rming motor vehicle emissions as the most likely dominant source of these pollutants. This pattern of concentrations is consistent with previously reported observations.

A laboratory - based comparison of PTR – MS measurements of benzene over three separate 1 hour periods with those obtained by GC – FID analysis of simultaneously collected adsorption trap samples showed good agreement. For the three periods, the PTR – MS recorded mean concentrations of 75 pptv, 72 pptv, and 78 pptv (30%), while the GC – FID technique measured concentrations of 69 pptv, 64 pptv, and 79 pptv (10%). All three values show an agreement of almost 90% or better, authenticating the concentrations observed at the monitoring station.


Figure 4.11 Benzene (closed circles) and NO (open circles) concentrations measured. (a) The average benzene concentration for 7 days; (b) a closer inspection of the data for 1 day only (Hayward et al ., 2002 ).

4.8.3 The Effect of Photocatalytic Oxidation on VOC Removal

Photocatalytic oxidation ( PCO ) is an innovative technique to remove VOCs (Tomp-kins, 2001 ; Yang et al ., 2007 ; Zhang et al ., 2007b ). The real - time monitoring system is a powerful tool in evaluating transient PCO air - cleaner performance. Besides,


it has been found that some harmful by - products may occur in the process of PCO of VOCs. Some by - products are not stable and react easily with other compounds. The production of these harmful by - products depends on the reaction conditions. It is diffi cult to study the transient air - cleaner ’ s performance and to reveal the harmful by - products without the real - time monitoring technique.

4.8.3.1 Detection of Harmful By - Product During the Removal of Toluene by PCO Glass - plate rectangular photocatalytic reactor is one of the simplest forms of PCO reactors to study the PCO performance of the given titania under various condi-tions. To evaluate the VOC removal effi ciency of the PCO reactor, the inlet and outlet VOC concentrations should be measured (Yang, Zhang and Zhao, 2004 ). In the research, INNOVA - 1312 was used to measure the VOC concentrations. The change of the gas concentrations after turning on and turning off the UV light was recorded by the real - time gas analyzer. From these results it is clear that the UV light is necessary for toluene removal by using PCO of titania because only when the UV light is on is the outlet toluene concentration much smaller than the inlet one (Figure 4.12 ). From the fi gure, it is also seen that (i) during the PCO process, the concentration of CO 2 is obviously higher than that in indoor air; (ii) the harmful by - product CO occurs due to the potocatalytic oxidation not being completed.

We recently built up an experimental system to further study the by - product of PCO of toluene (see Figure 4.13 ) (Mo, Xu and Zhang, 2008a ). A PTR – MS was applied to measure the inlet and outlet pollutant concentrations. Some results are shown in Figure 4.14 (Zhang et al ., 2008 ). It indicates that benzaldehyde, metha-nol, acetaldehyde, formic acid/ethanol, acetone/propionaldehyde and acetic acid

Figure 4.12 Schematic of experimental data for PCO reaction of toluene in a PCO reactor.


Figure 4.13 Schematic of the experimental system, including the gas preparation part and UV - PCO reaction cell.

Figure 4.14 Toluene and its intermediates conversions vs. time with inlet toluene concentration of 8 ppm, T = 24 – 26 ° C; RH = 47 – 50%; 254 nm: (a) UV intensity = 0.95 mW/cm 2 ; (b) UV intensity = 0.78 mW/cm 2 ; (c) UV intensity = 0.43 mW/cm 2 .


were the main intermediates in the gas phase. Under the experimental condition, acetaldehyde and acetone/propionaldehyde increased sharply at the beginning and trended to steady state. Simultaneously, the by - products of formaldehyde, metha-nol, formic acid/ethanol and benzaldehyde were generated but the increases were slower compared with acetaldehyde and acetone/propionaldehyde. For these, further research is needed.

4.8.3.2 Evaluating the Formaldehyde Removal Performance of PCO Reactors Applying fi ns coated with titania is an effective way to enhance the formaldehyde removal performance (Mo et al ., 2008b ). In order to compare the formaldehyde removal performance of annular photocatalytic reactors with and without the aforementioned fi ns, the PCO reactors without and with the fi ns operated in a sealed stainless steel chamber with formaldehyde in its air, respectively. Figure 4.15 shows the chamber formaldehyde concentration decay curves: (a) an annular PCO reactor without fi ns, (b) an annular PCO reactor with six fi ns. It can be seen that the reactor with fi ns is more effi cient than that without. The reason and the analysis are presented by Mo et al . (2008b) in detail.

4.8.4 Products of Ozone - Initiated Chemistry in a Simulated Aircraft Environment

Real - time monitoring like PTR – MS is a good choice when the component of the sample gas will be changed using a traditional sorbent or canister sampling method.

Figure 4.15 The chamber formaldehyde concentration decay curves: (a) an annular PCO reactor without fi ns, (b) an annular PCO reactor with six fi ns.


Wisthaler et al . (2005) used PTR – MS to examine the products formed when ozone reacted with the materials in a simulated aircraft cabin, including a loaded high - effi ciency particulate air ( HEPA ) fi lter in the return air system.

There can be both positive and negative artifacts when measuring the mixing ratios of aldehydes in ozone - containing air using conventional techniques such as sorbent or canister sampling. O 3 can react with unsaturated hydrocarbons on sorbents or in canisters to form aldehydes (positive artifacts), and it can react with carbonyl hydrazones, in the case of DNPH cartridges (negative artifacts). Given that PTR – MS analysis does not require preconcentration, a signifi cant source of artifacts can be ruled out. However, heterogeneous oxidation of reactive com-pounds in the inlet of the PTR – MS instrument may result in aldehyde artifacts. In the Wisthaler study, the entire inlet was made of clean prebaked Silcosteel tubing, with the exception of two Tefl on solenoid valves. In addition, the residence time in the inlet system was kept as short as possible (about 150 ms for air sampled upstream of the HEPA fi lter), to minimize ozone - inlet surface interactions. A similar inlet confi guration used in a previous study resulted in no detectable alde-hyde artifacts (detection limit of 20 pptv) at an ozone mixing ratio of 25 ppbv. Aldehyde artifacts, if present at all, are not expected to exceed 20 – 30 pptv for the individual species reported.

Four conditions were examined: cabin (baseline), cabin plus ozone, cabin plus soiled T - shirts (surrogates for human occupants), and cabin plus soiled T - shirts plus ozone. For each of the four simulated conditions, the mixing ratio of chemi-cals were continuously monitored in real time, using PTR – MS. Figure 4.16 shows the mixing ratios for ozone and the sum of the organic chemicals detected in the simulated cabin throughout these four days. In the absence of ozone generation, the ozone mixing ratio in the cabin air was quite low (typically < 5 ppbv). The ozone level in the cabin began to increase within minutes of turning on the ozone gen-erators. The PTR – MS signal began to increase at approximately the same time (see days 2 and 4). The ozone mixing ratio was smaller on day 4, when soiled T - shirts were present, than on day 2 when the cabin was empty. Conversely, the total PTR – MS signal reached a higher value on day 4 than on day 2.

The addition of ozone to the cabin without T - shirts, at concentrations typically encountered during commercial air travel, increased the mixing ratio (v:v concen-tration) of detected pollutants from 35 ppbv to 80 ppbv. Most of this increase was due to the production of saturated and unsaturated aldehydes and tentatively identifi ed low - molecular - weight carboxylic acids. The addition of soiled T - shirts, with no ozone present, increased the mixing ratio of pollutants in the cabin air only slightly, whereas the combination of soiled T - shirts and ozone increased the mixing ratio of detected pollutants to 110 ppbv, with more than 20 ppbv originating from squalene oxidation products (acetone, 4 - oxopentanal, and 6 - methyl - 5 - hepten - 2 - one). For the two conditions with ozone present, the more - abundant oxidation products included acetone/propanal (8 – 20 ppbv), formaldehyde (8 – 10 ppbv), nonanal (6 ppbv above), 4 - oxopentanal (3 – 7 ppbv), acetic acid (7 ppbv above), formic acid (3 ppbv above), and 6 - methyl - 5 - hepten - 2 - one (0.5 – 2.5 ppbv), as well as compounds tentatively identifi ed as acrolein (0.6 – 1 ppbv) and crotonaldehyde


(0.6 – 0.8 ppbv). The odor thresholds of certain products were exceeded. With an outdoor air exchange of 3 h − 1 and a recirculation rate of 20 h − 1 , the measured ozone surface removal rate constant was 6.3 h − 1 when T - shirts were not present, com-pared to 11.4 h − 1 when T - shirts were present.

4.9 Concluding Remarks

As mentioned at the beginning of this article, real - time monitoring of organic compounds is very useful for investigating indoor air quality both in the fi eld and in laboratories. Much research in the fi eld of indoor air quality relies on these techniques. It is however crucial to apply the techniques with an awareness of their limitations and the relative errors of the measured data.

Real - time monitoring techniques for organic compounds encompass many dis-ciplines. Many problems have been solved, and yet more need further investiga-tion. With the development of related science and technologies, real - time monitoring techniques will continue to develop. There are still many challenges and opportunities to researchers in the related research areas.

Figure 4.16 Total mixing ratio (in units of ppbv) recorded with PTR – MS instrument from day 1 through day 4. The ozone mixing ratio (in units of ppbv) during this period is also shown.

References 97

Acknowledgments

This work was supported by National Nature Science Foundation of China (The grant Nos. are 50436040 and 50478012) and the National Eleventh Five - year Plan Project of China (No. 2006BAJ02A08). The authors wish to express heartfelt thanks to Qiujian Xu for his help on writing this article and to Susan for her help on revising this article.

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Mo , J.H. , Zhang , Y.P. , Yang , R. et al . ( 2008b ) Infl uence of fi ns on formaldehyde removal in annular photocatalytic reactors . Building and Environment , 43 , 238 – 45 .

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Park , J.H. , Spengler , J.D. , Yoon , D.W. et al . ( 1998 ) Measurement of air exchange rate of stationary vehicles and estimation of in - vehicle exposure . Journal of Exposure Analysis and Environmental Epidemiology , 8 ( 1 ), 65 – 78 .

Poschl , U. , Williams , J. , Hoor , P. et al . ( 2001 ) High acetone concentrations throughout the 0 – 12 km altitude range over the tropical rainforest in Surinam . Journal of Atmospheric Chemistry , 38 ( 2 ), 115 – 32 .

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Schmid , W. , Barsan , N. and Weimar , U. ( 2003 ) Sensing of hydrocarbons with tin oxide sensors: possible reaction path as revealed by consumption measurements . Sensors and Actuators B - Chemical , 89 ( 3 ), 232 – 6 .

Shin , H.S. , Lee , J.K. , Ahn , Y.C. et al . ( 2005 ) Measurement of indoor air quality for ventilation with the existence of occupants in schools . Journal of Mechanical Science and Technology , 19 ( 4 ), 1001 – 5 .

Stefaniak , A.B. , Breysse , P.N. , Murray , M.P.M. et al . ( 2000 ) An evaluation of employee exposure to volatile organic compounds in three photocopy centers . Environmental Research , 83 ( 2 ), 162 – 73 .

Tompkins , D.T. ( 2001 ) Evaluation of photocatalytic air cleaning capability: a literature review and engineering analysis . ASHARE Research Project RP - 1134.

Warneke , C. , Holzinger , R. , Hansel , A. et al . ( 2001 ) Isoprene and its oxidation products

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Environmental Test Chambers and Cells Tunga Salthammer

5

5.1 Introduction

Evaluation of VOC and SVOC emission potential of individual products and mate-rials under indoor - related conditions and over defi ned timescales requires the use of climate - controlled emission testing systems, so - called emission test chambers and cells, the size of which can vary between a few cm 3 and several m 3 , depending on the application. In Figure 5.1 the dots ( • ) represent volumes of test devices reported in the literature. From this size distribution they can be classifi ed as large scale chambers, small scale chambers, micro scale chambers and cells. The selec-tion of the systems, the sampling preparation and the test performance all depend on the task to be performed. According to ISO, chambers and cells are defi ned as follows:

• Emission test chamber: Enclosure with controlled operational parameters for the determination of volatile organic compounds emitted from building products (ISO, 2006a ).

• Emission test cell: Portable device for the determination of volatile organic compounds emitted from building products. The emission cell is placed against the surface of the test specimen, which thus becomes part of the emission cell (ISO, 2006b ).

During an emission investigation the product/material to be investigated is tested with regard to temperature ( T ), relative humidity (r.h.), air exchange rate ( n ), air velocity and product loading factor ( L = ratio of surface of product to be investi-gated to the volume of the emission test chamber) under standardized conditions in the testing device that can be sealed gas tight against the outside atmosphere. The test procedure is suitable for emission investigations of both surfaces and of volume samples. This is a convention process where the boundary conditions are selected in such a way that they refl ect those to be found in realistic indoor rooms. When interpreting the results of test chamber/cell investigations, under certain


101

102 5 Environmental Test Chambers and Cells

circumstances, the limiting factor must be taken into account that not all realistic conditions to be found in an indoor room can be simulated.

Emission measurements are performed for the following purposes as for example:

• compiling of substance - specifi c emission data from various sources to back up fi eld investigations into indoor air quality;

• determination of the infl uence of environmentally relevant factors such as temperature, humidity and air exchange on the emission characteristics of the products;

• processing of characteristic data to estimate product emissions;

• processing of characteristic emission data to develop models that can be used to predict indoor concentrations;

• determination of sorption and diffusion properties of products;

• ranking of various products and product types on the basis of the characteristic emission data.

5.2 Characteristics of Chambers and Cells

The fi rst room - sized test chambers were developed in the mid - 1970s in the course of the introduction of building authority regulations concerning formaldehyde emissions from wood particle boards. Interest soon expanded to include VOCs and nowadays chamber investigations are probably the most important method used for the determination of the emission characteristics of materials. Moreover, a reduction in the chamber size had to be achieved for practical reasons. Today, emission test chambers typically range in size from 50 m 3 (Singer et al ., 2006 ) to 0.020 m 3 (Salthammer et al ., 1995 ) (see Figure 5.1 ). A common interior volume for an emission test chamber is 1 m 3 and the interior is usually made of glass (Salthammer, 1997 ) or stainless steel (Meyer et al ., 1994 ) (see Figure 5.2 ).

It is often desired to have a device for a quality control of products with regard to emissions arising during the utilization phase. To enable action to be taken

Figure 5.1 Classifi cation of test chambers and cells.

5.2 Characteristics of Chambers and Cells 103

during the production process as quickly as possible, fast emission investigations accompanying the production process would be a worthwhile objective. It is there-fore desirable to have a measuring system that can be used to carry out emission testing and quality assurance on location. The relevant principle of a transportable emission testing cell for mobile application was implemented in Scandinavia for the fi rst time in 1991 with the so - called FLEC (see Figure 5.2 ) (Wolkoff, 1996 ; Wolkoff, Salthammer and Woolfenden, 2005 ). The FLEC opens up the opportunity of carrying out nondestructive emission testing on surfaces within the framework of fi eld investigations. In this way it is possible to identify emissions from building products when they are already installed and also the sources for air - polluting substances.

A so - called Micro Chamber ( μ - CTE) (see Figure 5.2 ) with an interior volume of approximately 45 ml has been designed by Markes Int. (Schripp et al ., 2007 ). The μ - CTE comprises six individual stainless steel cylindrical chambers ( d = 4.5 cm), located in one unit. All six chambers are supplied simultaneously with the same, controlled fl ow of synthetic air. By reducing the chamber volume, it was intended to reduce typical emission test times but still generate meaningful emission data – i.e. results that correlate with data from conventional emissions test cham-bers. The μ - CTE was not intended to replace standard emission test facilities. In fact it is a complementary tool intended to produce fast information about the composition and level of VOC emissions for development of new, low - emission products/materials. Micro chamber measurements have already been shown to

Figure 5.2 Different types of devices for emission testing: A: Field and Laboratory Emission Cell (FLEC); B: WKI 1 m 3 glass chamber with Fast Mobility Particle Sizer (FMPS); C: 1 m 3 stainless steel chamber; D: Microchamber; E: WKI 48 m 3 stainless steel chamber.


provide a useful basis for fast emission screening prior to formal product certifi ca-tion. Moreover, since the device has only a small chamber volume and offers a relatively high loading factor, it has also been shown to offer enhanced sensitivity relative to conventional chambers for compounds of low volatility (SVOCs) (Scherer et al ., 2006 ) which is due to reduced sink effects.

Emission test chambers, cells and analytical procedures are now standardized by ISO, ASTM and other authorities (see Table 5.1 ). However, it is interesting to note that only the ASTM standards take sink effects (see Section 5.3 ) into account. Different types of testing facilities have different properties regarding amount of test specimen, dynamics, time, result and cost. In a small device the test is gener-ally time effi cient, but on the other hand sample inhomogeneities will signifi cantly infl uence the results. In Figure 5.3 some trends are shown in dependency of the chamber/cell size.

Table 5.1 Standard chamber and cell methods for emission testing.

Standard Description

ISO 16000 - 6 (2004) Determination of VOCs in test chamber air ISO 16000 - 9 (2006a) Emission test chamber method ISO 16000 - 10 (2006b) Emission test cell method ISO 16000 - 11 (2006c) Preparation of test specimens ISO 16000 - 25 (under development) (2008) Micro - chamber method ASTM D 5116 (2006) Small scale chamber ASTM D 6670 (2001) Large scale chamber ASTM D7143 (2005) Emission cell

Figure 5.3 Infl uence of the chamber size on test parameters and results.

5.3 Sink Effects 105

5.3 Sink Effects

A sink effect is the fact that the released components partially adsorb within the test chamber, for example, at the chamber walls (Sollinger, Levsen and W ü nsch, 1993 ). This can result in an incorrect reading of the concentration determined at the chamber outlet, which can lead to the wrong emission rate being computed. In principle every test chamber demonstrates a low sink effect. The degree of adsorption on chamber walls and in materials as well as the extent of the recovery can vary extremely for different VOCs (Meininghaus, Salthammer and Kn ö ppel, 1999 ). This depends in part on the volatility of the relevant substance. It was recently demonstrated that adsorption within the test chamber is favored as the boiling point increases (Uhde and Salthammer 2006 ). A comparatively higher polarity of a compound can also be favorable to the adsorption at the chamber walls. Sollinger, Levsen and W ü nsch (1993) also showed that the sink effect is clearly increased by introducing an adsorbing sample surface (carpet) into the test chamber. The fact that the sample to be investigated in a chamber also acts as a sink has been reported by several authors (Colombo et al ., 1993 ; Tichenor, 1996 ; Jorgensen, Bjorseth and Malvik, 1999 ). An important factor infl uencing sinks is the air distribution in the device, which can be determined by use of anemometers (Salthammer et al ., 1995 ; Uhde, Borgschulte and Salthammer, 1998 ) or can be calculated by computational fl uid dynamics (Schripp et al ., 2007 ).

While the sink effects caused by the chamber itself can be reduced to a minimum by using appropriate construction materials, it is not possible to infl uence the sink effects attributable to the actual sample. Many materials subjected to emission testing are also good sinks for the substances emitted by them. Especially with porous building materials with a large surface area or with foams it must be expected that the intrinsic material sink effect leads to a clear delay of the sub-stance emission. An experiment was performed where the test substance 1 - bromo - 3 - chloro - benzene (boiling point = 196 ° C) was dosed into a 1 m 3 glass chamber continuously until a constant concentration was achieved (sorption phase). Then the dosing was stopped and the decay of concentration was observed (desorption phase). The experiment was then repeated using a gypsum board in the chamber. In Figure 5.4 the concentration decay curves with and without sink are compared. The dotted curve represents the theoretical decay as being calculated from the air exchange rate k 2 = 1.0 h − 1 . It is obvious that the decay of the concentration of the test substance was delayed in both cases. The empty chamber already provides a small sink for 1 - bromo - 3 - chloro - benzene, but the effect was much stronger in the presence of gypsum board. The sink effect of the sample material therefore means that the measured emission rate is smaller than the actual rate, as emitted sub-stances are initially absorbed in the sink. At the same time the measured duration of the emission process is longer, as even after the completion of the actual emis-sion substances are still released from the sink and lead to the chamber concentra-tion dropping more slowly. This sink effect caused by the sample to be investigated plays a role in each chamber investigation and cannot be avoided. For this reason


it is of fundamental importance to guarantee that the chamber has as good a recovery as possible. Uhde and Salthammer (2006) have evaluated different test chambers by statistical cluster analysis. Principal component analysis ( PCA ) revealed that the sink effect essentially depends on a compound ’ s boiling point.

In connection with the sink effect, it should also be taken into consideration that a so - called ‘ memory effect ’ can result for subsequent investigations, which must be avoided by appropriate removal of contaminants. It is advantageous here if the test chamber can be thoroughly heated through at higher temperatures for the purpose of thermal cleaning.

5.4 Calculation of Emission Rates

Air measurement in a chamber or cell initially produces the concentration C (t) at the time t of the measurement. To enable better comparability of the measured data the specifi c emission rate ( SER ) independent of air exchange and loading is to be preferred. The SER describes the product - specifi c emission behavior, for example, as area - specifi c emission rate (SER A ) with the unit μ g/(m 2 h) or as unit - based specifi c emission rate (SER U ) with the unit μ g/(unit h).

The time - dependent determination of the emission potential is carried out according to the balance Equation 5.1 , whereby C (t) = chamber concentration in μ g/m 3 , n = air exchange (in h − 1 ) and L = loading in (m 2 /m 3 ).

d d SER t tC t L n C= ( ) − ( ) (5.1)

Figure 5.4 Decay curves of 1 - bromo - 3 - chloro - benzene in the 1 m 3 glass chamber: ( � ) with sink (gypsum board); ( • ) without sink; ( · · · · · ) theoretical decay (Uhde and Salthammer, 2006 ).

5.4 Calculation of Emission Rates 107

For a decaying concentration - time function SER(t) is obtained from (5.1) by transi-tion to the difference quotient according to Equation 5.2 .

SER t t( ) = ( ) + ( )[ ]Δ ΔC t n C L (5.2)

with

Δ ΔC t C C t t C C t ti i i i i i i i i i= −( ) −( ) + −( ) −( )[ ]− − + +1 1 1 1 2 (5.3)

Thus, if there are n + 1 experimental data available for concentration, n − 1 emis-sion rate values can be obtained by this method. In the steady state (d C /d t = 0) Equation 5.2 progresses to Equation 5.4 .

SER = ( )n C L (5.4)

In the simplest case a compound is emitted from a sample into the chamber and removed with exhaust air. This so - called ‘ dilution - model ’ has been introduced by Dunn and Tichenor (1988) for a source with constant and exponentially decaying emission rate, respectively. A more sophisticated model requires the compart-ments source, chamber, exhaust and sink. The rate constants k i describe the exchange of mass between different sections. The parameter k 2 represents the air exchange rate n . Figure 5.5 shows a scheme of the two - sink model after De Bortoli et al . (1996) as an extension of the ‘ full - model ’ ansatz after Dunn and Tichenor (1988) with k 1 − 5 ≠ 0 und k 6 = 0. Equations 5.5 and 5.6 are solutions of the simple ‘ dilution - model ’ with k 3 − 6 = 0 for constant emission and exponentially decaying emission, respectively.

Ck

k Vet k t( ) = −( )−1

2

1 2 (5.5)

Ck

k k Ve et

SER k t k t( ) =−( )

−( )− −1

1 2

2 1 (5.6)

Figure 5.5 Kinetic scheme of the two - sink model (De Bortoli et al ., 1996 ).


Equations 5.7 and 5.8 are simplifi ed forms using the general rate constants b 1 , b 2 and b 3 of the more complex solutions with k 3,4 ≠ 0 und k 5,6 = 0 for constant emis-sion and exponentially decaying emission, respectively. Equations 5.7 and 5.8 have also been applied for interpolation purposes.

C a e a et b t b t( ) = −( ) − −( )− −1 21 11 2 (5.7)

C a e a e a et b t b t b t( ) = + +− − −1 2 3

1 2 3 (5.8)

The complete solution with k 1 − 6 ≠ 0 is described by De Bortoli et al . (1996) . The properties of different models were compared by Colombo and Bortoli (1992) . However, the calculation of emission rates using physically based models requires nonlinear curve fi tting and a suffi cient number of data points. Large errors in parameter estimates can result from rough chamber data and/or wrong models (Salthammer, 1996 ).

5.5 Kinetics and Mass Transfer

The dynamics of emission processes from building product surfaces have been studied in detail and a number of physical and empirical models of different complexity have been described (Tichenor, 1996 ). The emission can be character-ized by two fundamental physical processes (see Figure 5.6 ) (Sparks et al ., 1996 ):

i) Gas - phase mass transfer (i.e., external diffusion); ii) Source - phase mass transfer (i.e., internal diffusion).

The gas - phase mass transfer model (i) is based on molecular diffusion across a laminar boundary layer as described in Equation 5.9 .

Figure 5.6 Diagram of the kinetic processes involved in the mass transfer model of VOCs from material surfaces; adopted from (Sparks et al ., 1996 ).

5.6 Application of Test Chambers and Cells 109

SER S i g S i= ⋅ −( ) = ⋅ −( )DC C k C C

δ (5.9)

SER is the specifi c emission rate, D is the diffusion coeffi cient, δ is the thickness of the boundary layer, C S is the concentration of the target VOC at the source surface and C i is the concentration of the target VOC in the air and k g is the gas - phase mass transfer coeffi cient.

Process (ii) is limited by diffusion from the interior of the source to the surface and can be described by Equation 5.10 .

SER s S i= ⋅ −( )k m m (5.10)

Here, m s is the mass of the target VOC in the source, m i is the mass of the target VOC at the surface and k s is the source - phase mass transfer coeffi cient.

In terms of the comparability of emission test results obtained from different chambers and cells, three different scenarios have to be considered. For k g >> k s the emission is controlled by the external diffusion process and the thickness of the boundary layer δ is directly related to the air velocity above the surface. This applies to most wet - applied or liquid products during the drying/curing phase. In this case, the air fl ow conditions in the test facility (i.e., air velocity and turbulence) are important. This means that precise control of the air velocity may be critical in the short term (i.e., 1 – 14 days) if it is not to infl uence the emission test result. For k g << k s the emission is controlled by the internal diffusion process and the infl uence of the air fl ow condition in the test facility should be negligible. This applies to most materials manufactured in the solid phase and to wet - applied or liquid products after they have dried or cured. A more diffi cult situation appears for k g ≈ k s or if the ratio k s / k g changes over time. This situation arises in case of curing two - component diisocyanate adhesives, where a two - step process can be observed. In the fi rst step, the emission is dominated by surface evaporation, and the decay of emission is mainly caused by the decrease in monomer content due to reaction. In the second step, the release is limited by internal diffusion (Wirts et al ., 2003 ). Little and colleagues have successfully applied mass transfer models for describing the emission of plasticizers from PVC fl oorings (Cox, Little and Hodgson, 2002 ; Xu and Little, 2006 ).

5.6 Application of Test Chambers and Cells

Standardized Emission Testing This is probably the most frequent application of test chambers and cells, because indoor related materials and products need to be evaluated for the release of volatile chemicals in order to ensure a healthy indoor climate. Many procedures have been established for different types of products. A very well - known scheme was developed by the German ‘ Committee for Health - related Evaluation of Building Products ’ (AgBB) for the evaluation of building


products on the basis of their emission properties (AgBB, 2008 ). The product is tested in an emission test chamber or an emission test cell over a period of 28 days. It has to meet seven separate requirements regarding single substances, VOCs, TVOC and ∑ SVOC at different times to be ranked as ‘ suitable for indoor use ’ . AgBB has planned to complement this procedure by sensory testing in the near future (see also Chapter 6 ).

Fogging This test is a special case of test chamber examination – SVOCs such as plasticizers and fl ame retardants are precipitated inside the chamber on a cooled surface. The method had initially been developed for examining automotive parts in order to determine the portion of ‘ fogging - active ’ substances. It can also be used for examining other products used indoors. The fogging value (in μ g), usually determined over a 14 - day period, is a characteristic for the SVOC quantity which can be expected to condense on cold indoor surfaces. The fogging method is based on a convention. If results are to be compared, the studies must be carried out in an identical manner (Uhde et al ., 2001 ; Wensing, Uhde and Salthammer, 2005 ).

Combination of Chambers and Cells Meininghaus and Uhde (2002) have combined chamber and FLEC (so - called one - fl ow system) to observe the diffusion of a VOC mixture through a test sample. A 23 l glass chamber (operated under static condi-tions) was used as a limited reservoir for gaseous VOCs. To start the experiment, the chamber lid was replaced by the test specimen and the FLEC (operated under dynamic conditions) was placed on the outer surface (see Figure 5.7 ). The principle of the so - called two - fl ow system setup is also depicted in Figure 5.7 (Meininghaus, Salthammer and Kn ö ppel, 1999 ). Two FLECs are installed face - to - face, separated by the test specimen. The air supply for FLEC1 passes through a VOC vapor gen-eration system (Meininghaus, Schauenburg and Kn ö ppel, 1998 ). The air is then directly led into the cell. Concentrations of supply air (inlet) to FLEC1 and exhaust

Figure 5.7 Combination of 23 l chamber with FLEC (one fl ow system) (Meininghaus and Uhde, 2002 ) and combination of two FLECs (two fl ow system with 1: supply air; 2: mass fl ow controller; 3: VOC generator; 4: FLEC1; 5: FLEC2; 6: test specimen; 7: valve; A, B, C: sampling ports) (Meininghaus et al ., 1999 ).

5.6 Application of Test Chambers and Cells 111

air (outlet) of both FLECs were measured. A similar device for diffusion testing using CLIMPAQs has been described by Meininghaus, Gunnarsen and Knudsen ( 2000 ). Clausen et al . (2004) have used the FLEC face - to - face setup for the inves-tigation of DEHP emission from PVC. Other scenarios describe the application of chambers and cells connected in series for the investigation of SVOC emissions (see Table 5.2 ).

Indoor Chemistry Various terpenes and terpenoids are emitted from household products and building materials. Ozone that has entered from outdoors or has been generated indoors can react with these compounds, either in the gas phase or on the surface of materials. The resulting oxidation products will contribute to the production and growth of meaningful quantities of secondary organic aerosol s ( SOA ). The formation and growth of SOA can be studied under controlled condi-tions in test chambers (see also Chapter 13 ).

Large Chamber Studies Real - life scenarios can be investigated using large (walk - in - type) chambers. As an example, researchers from the Technical University of

Table 5.2 Examples of advanced applications of test chambers and cells.

Application Reference

23 l chamber + FLEC Meininghaus and Uhde (2002)

FLEC + FLEC (sandwich like or double FLEC) Meininghaus, Salthammer and Kn ö ppel (1999) Clausen et al . (2004)

CLIMPAQ + CLIMPAQ Meininghaus, Gunnarsen and Knudsen (2000)

23 l chamber + 23 l chamber Jann and Wilke (1999)

FLEC + FLEC (connected in series) Clausen et al . (2004)

Chamber + fogging sampling Uhde et al . (2001) Wensing, Uhde and Salthammer (2005)

Indoor chemistry studies in chamber Sarwar et al . (2003) Destaillats et al . (2006) Coleman et al . (2008) Toftum et al . (2008)

Large chamber studies Bak ò - Bir ò et al . (2004) Tamas et al . (2006) Salthammer and Mentese (2008)

Electronic equipment studies in chamber Schripp et al . (2008) Wensing et al . (2008)


Denmark have studied the infl uence of ventilation, offi ce equipment and terpene/ozone reactions on perceived air quality. Salthammer and Mentese (2008) have studied formaldehyde and VOC concentrations in a test house using the WKI 48 m 3 stainless steel chamber (see Figure 5.8 ).

Testing of Electronic Offi ce Equipment This is a relatively new application of chamber testing and mainly applied to laser printers and hardcopy devices. In contrast to the testing of building products the available time frame is relatively short because a printing event normally proceeds within ten minutes. VOCs, ozone and particles are typically measured (see also Chapter 17 ).

5.7 Final Remarks

For many years test chambers and cells belong to the most important tools for the simulation of indoor related conditions and for the evaluation of emission rates. It can be assumed that their application will become even more important in the near future. Moreover, powerful kinetic models are now available that help to understand emission characteristics of sources. The current trend to develop devices en miniature brings us to the borderline between emission and content analysis. It will be interesting to see new chamber designs and intelligent applica-tions for indoor related studies.

Figure 5.8 Test house study in the WKI 48 m 3 stainless steel chamber (Salthammer and Mentese, 2008 ).

References 113

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Destaillats , H. , Lunden , M.M. , Singer , B.C. , Coleman , B.K. , Hodgson , A.T. , Weschler , C.J. and Nazaroff , W.W. ( 2006 ) Indoor sec-ondary pollutants from household product emissions in the presence of ozone: a bench - scale chamber study . Environmental Science and Technology , 40 , 4421 – 8 .

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ISO ( 2006c ) 16000 - 11 . Indoor Air – Determination of the Emission of Volatile Organic Compounds from Building Products and Furnishing – Sampling, Storage of Samples and Preparation of Test Specimens ,


International Organization for Standardization , Geneva, Switzerland .

ISO ( 2008 ) 16000 - 25 (under development). Indoor Air – Determination of the Emission of Semi - Volatile Organic Compounds by Building Products – Micro - Chamber Method , International Organization for Standardization , Geneva, Switzerland .

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Investigation Concepts and Quality Guidelines


Part Two

Standardized Methods for Testing Emissions of Organic Vapors from Building Products to Indoor Air Elizabeth Woolfenden

6

6.1 Introduction: The Need for Standardization

Testing volatile organic emissions to air from construction materials and con-sumer products used indoors is the subject of intensive standard method develop-ment at the current time. Most of the world ’ s leading national and international standards agencies are involved to a greater or lesser extent. There are several reasons for this sudden burst of activity. Top of the list are recent regulatory developments such as:

• The EU Action Plan on Environment and Health ( EHAP ) 2004 – 2010 (COM (2004) 416 fi nal) which included ‘ improving indoor air quality ’ as one of its action points. This led to:

� the 2005 European Parliament resolution on indoor air quality (Ries, 2005 ) which called for both research into the impact of new construction materials on human health and for cross - border chemical emissions labeling schemes for construction products.

� The inclusion of indoor air quality and associated health impacts within the Seventh EC ‘ Framework Program of Research ’ (FP7, 2007 – 2013) ( http://cordis.europa.eu/fp7/home_en.html ).

• The 1989 EC Construction Products Directive ( CPD ) (1989/106/EEC) which includes assessment of chemical emissions as one of six ‘ Essential Requirements ’ ( ER s). The CPD is currently under revision and is the subject of EC mandate M/366 (2005) .

• The 2006 European ‘ REACH ’ regulation (Registration, Evaluation, Authorisation and restriction of Chemicals) (2006/121/EC) which covers intentional and unintentional release of chemicals from ‘ articles ’ and preparations (including many consumer goods).


119

120 6 Standardized Methods for Testing Emissions of Organic Vapors

• The 2008 Californian ‘ Green Chemistry ’ bill: Assembly Bill 1879 – nicknamed ‘ Californian REACH ’ .

• The EC directive on Energy Performance of Buildings (EPBD) (2002/91/EC) which stresses the importance of maintaining indoor air quality as buildings are sealed to minimize energy loss.

• German national requirements for the certifi cation of fl oor covering materials.

The World Health Organization ( WHO ) also began an initiative in 2006 to develop guidelines for indoor air quality ( IAQ ) (WHO, 2006 ), recognizing the importance of IAQ as a key risk factor for human health in both developing and developed countries. WHO refer to the many potentially hazardous compounds released indoors due to; combustion (e.g., coal fi res and smoking), emissions from prod-ucts used indoors (construction products, furnishings, consumer goods, cleaning products, etc.) and microbial pollution (bacteria and molds).

A plethora of product labeling schemes has grown out of this regulatory activity each specifying its own chemical emissions testing methodology. Relevant exam-ples include: the German AgBB scheme (AgBB, 2001/2008 ); the US Greenguard Certifi cation Program ( www.greenguard.org ); the Californian Collaborative for High Performance Schools, 2004 ; the Finnish M1 label ; the GUT carpet label (GUT, 2008 ) and the EMICODE label for fl ooring installation products (GEV EMICODE, 2001 ).

Public awareness of air quality as a ‘ green ’ issue and associated health concerns have also boosted interest in this area. The consumer has begun to demand the right to make an ‘ informed choice ’ and select products on the basis of chemical emissions. As public awareness of this issue has grown, companies ranging from car manufacturers to paint producers have started to use ‘ low emissions ’ as a marketing tool, that is, as a factor by which to differentiate themselves and their products from their competitors.

To satisfactorily meet and address these regulatory and market needs, standard methods for emissions testing need to be both robust and repeatable for competent laboratories to carry out, and practical and affordable for manu-facturing industry. Uniformity of test methods, between countries and markets is also important if manufacturers are to avoid having to submit the same products for emissions testing by multiple different emissions certifi cation protocols.

6.2 Materials Emissions Testing: A Challenge for Method Standardization

Testing very volatile, volatile and semi - volatile organic chemical (VVOC, VOC and SVOC) emissions from diverse product types to satisfy the differing needs of various national and international regulations and consumer or market inter-ests, is a diffi cult thing to do. Emission testing is best carried out under conditions

6.2 Materials Emissions Testing: A Challenge for Method Standardization 121

that are as near to ‘ real - world ’ as practically possible in order to ensure that test data relate as closely as possible to the potential impact a product might have on indoor air quality. Among the main challenges facing prospective emissions analysts are:

• the number and diversity of product types affected; • the number and diversity of potential target compounds; • the limited toxicological data available for most chemicals with respect to

inhalation exposure; • minimizing the introduction of uncertainty (error) in the multistep materials

emissions testing process; • the number and nonuniformity of emissions - related certifi cation protocols and

product standards that are currently in circulation or development.

6.2.1 The Range of Products and Materials Requiring Emissions Testing

Even within the confi nes of construction products and furnishings, materials ranging from textiles to paints and from concrete leveling compounds to vinyl fl ooring all present widely different sample matrices and preparation require-ments for emissions measurement. Each of these samples has to be prepared for emissions testing in a way that is as close as possible to real - world application of the product during building construction and decoration. This means that several different sample preparation protocols are required. While it is perfectly possible to specify individual preparation procedures/protocols for each different product type (for example see EN/ISO 16000 - 11 in Table 6.1 ), the requirement for multiple procedures nevertheless makes life more diffi cult for test laboratories as they try to ensure that test data aren ’ t compromised by variability during the sample prepa-ration stage.

6.2.2 The Range of Potential Target Compounds

The challenge posed by the range and number of compounds of interest is even greater than that caused by the different material and product types. The category ‘ vapor - phase organic compounds ’ theoretically includes over a million different species which in turn cover an enormous range of chemistries, volatilities and toxicities. While sorbent tube sampling with TD and GC – MS is compatible with the widest range of vapor - phase organic chemicals and associated concentration levels, no single analytical method can be used to measure all such compounds. Some recent emissions test protocols or methods have therefore taken steps to restrict the range of target compounds, either to those with known higher risk of health effects or to those compounds that are known to emit from a given category of material or product. Nevertheless, the sheer number of potential target analytes adds to the complexity of the emissions testing process.


Table 6.1 List of prominent standard material emissions test methods and protocols.

Responsible agency Standard number Method title Comments

ISO ISO 16000 - 3 Indoor air – Part 3: Determination of formaldehyde and other carbonyl compounds – Active sampling method

ISO ISO 16000 - 6 Indoor air – Part 6: Determination of VOCs in indoor and test chamber air by active sampling on Tenax TA sorbent, thermal desorption and gas chromatography using MS/FID

CEN/ISO EN/ISO 16000 - 9 Indoor Air – Part 9: Emission test chamber method

Will supercede EN 13419 - 1

CEN/ISO EN/ISO 16000 - 10 Indoor Air – Part 10: Emission test cell method


CEN/ISO EN/ISO 16000 - 11 Indoor Air – Part 11: Procedure for sampling and storage of samples and preparation of test specimens


ISO ISO DIS 16000 - 23 Performance test for evaluating the reduction of formaldehyde concentrations by sorptive building materials

ISO ISO DIS 16000 - 24 Performance test for evaluating the reduction of volatile organic compounds and carbonyl compounds without formaldehyde concentrations by sorptive building materials

ISO ISO DIS 16000 - 25 Determination of the emission of semi - volatile organic compounds by building products – Micro - chamber method



CEN/ISO EN/ISO 16017 - 1 Air quality – Sampling and analysis of VOCs in ambient, indoor and workplace air (and small and large - scale materials emission chambers) by sorbent tube/thermal desorption/capillary gas chromatography. Part 1: Pumped sampling.

Primary EN/ISO standard for monitoring vapor - phase VOCs with pumped sorbent tubes and thermal desorption – GC( - MS) analysis

ISO ISO DIS 10580 Resilient, textile and laminate fl oor coverings. Test methods for emissions of volatile organic compounds

ISO ISO WD 12219 - 1 Road vehicles – Part 1: Whole vehicle test chamber. Specifi cation and method for the determination of volatile organic compounds in car interiors

CEN EN 717 - 1 Wood - based panels – Determination of formaldehyde release. Part 1: Formaldehyde emission by the chamber method

CEN EN 717 - 2 Wood - based panels – Determination of formaldehyde release. Part 2: Formaldehyde release by the gas analysis method

CEN EN 717 - 3 Wood - based panels – Determination of formaldehyde release. Part 3: Formaldehyde release by the fl ask method

CEN ENV 1250 - 1 Wood preservatives – Methods for measuring losses of active ingredients and other preservative ingredients from treated timber – Part 1 Laboratory method for obtaining samples for analysis to measure losses by evaporation to air

End - sealed preservative treated and untreated timber test specimens are prepared and placed in a chamber through which pure air is passed at a controlled rate. The air leaving the chamber is sampled intermittently

Table 6.1 Continued



CEN EN 71 - 11 Safety of toys – Part 11 – Organic chemical compounds methods of analysis

CEN EN 13999 - 1 Adhesives – Short term method for measuring the emissions properties of low - solvent or solvent - free adhesives after application. Part 1: General procedure.

CEN EN 13999 - 2 Adhesives – Short term method for measuring the emissions properties of low - solvent or solvent - free adhesives after application. Part 2: Determination of VOCs.

CEN EN 13999 - 3 Adhesives – Short term method for measuring the emissions properties of low - solvent or solvent - free adhesives after application. Part 3: Determination of volatile aldehydes.

CEN EN 13999 - 4 Adhesives – Short term method for measuring the emissions properties of low - solvent or solvent - free adhesives after application. Part 4: Determination of volatile diisocyanates

CEN TC351 Working Draft

Construction products - Assessment of emissions of regulated dangerous substances from construction products - Determination of emissions into indoor air

Table 6.1 Continued



ASTM D 5116 Standard guide for small - scale environmental chamber determinations of organic emissions from indoor material/products

Primary ASTM method for emissions testing using small chambers

ASTM D 5172 Test method for determination of formaldehyde and other carbonyl compounds in air (active sampler methodology)

ASTM D 6196 Standard practice for selection of sorbents, sampling and thermal desorption analysis procedures for VOCs in air (and material emissions chambers)

Primary ASTM standard for monitoring vapor - phase VOCs with sorbent tubes and thermal desorption – GC( - MS) analysis (pumped and diffusive)

ASTM D 6670 Standard practice for full - scale chamber determination of VOCs from indoor materials/products

ASTM D 7143 Standard practice for emission cells for the determination of VOCs from indoor materials/products

ASTM D 6330 Standard practice for the determination of VOCs (excluding formaldehyde) emissions from wood - based panels using small environmental chambers under defi ned test conditions

ASTM D 6803 Standard practice for testing and sampling of VOCs (including carbonyl compounds) emitted from paint using small environmental chambers

Table 6.1 Continued



ASTM D 6177 Standard practice for determining emissions profi les of VOCs emitted from bedding sets

ASTM E 1333 Standard test method for determining formaldehyde concentrations in air and emission rates from wood products using a large chamber

ASTM D 5582 Standard test method for determining formaldehyde levels from wood products using a dessicator

ASTM D7339 Test Method for Determination of Volatile Organic Compounds Emitted from Carpet Using a Specifi c Sorbent Tube and Thermal Desorption/Gas Chromatography

JIS A 1460 2001 Building boards Determination of formaldehyde emission – Desicator method

JIS A 1901 2003 Determination of the emission of volatile organic compounds and aldehydes for building products – Small chamber method

MOD: EN/ISO 16000 - 9

JIS A 1902 - 1 2006 Determination of the emission of volatile organic compounds and aldehydes for building products – Sampling, preparation of test specimens and testing condition – Part 1: Boards, wallpaper and fl oor materials

MOD: ISO 16000 - 10

Table 6.1 Continued



JIS A 1902 - 2 2006 Determination of the emission of volatile organic compounds and aldehydes for building products – Sampling, preparation of test specimens and testing condition – Part 2: Adhesives

MOD: ISO 16000 - 10

JIS A 1902 - 3 2006 Determination of the emission of volatile organic compounds and aldehydes for building products – Sampling, preparation of test specimens and testing condition – Part 3: Paints and coating materials

MOD: ISO 16000 - 10

JIS A 1902 - 4 2006 Determination of the emission of volatile organic compounds and aldehydes for building products – Sampling, preparation of test specimens and testing condition – Part 4: Heat - Insulating material boards

MOD: ISO 16000 - 10

JIS A 1903 2008 Determination of the emission of volatile organic compounds (VOC) for building products – Passive method

JIS A 1904 2008 Determination of the emission of semi volatile organic compounds for building products – Micro chamber method

proposed as ISO work item

JIS A 1905 - 1 2007 Performance test of sorptive building materials of reducing indoor air pollution with small chamber – Part 1: Measurement of adsorption fl ux with supplying constant concentration of Formaldehyde


Table 6.1 Continued



JIS A 1905 - 2 2007 Performance test of sorptive building materials reducing indoor air pollution with small chamber – Part 2: Measurement of capability for suppressing formaldehyde emission

JIS A 1906 2008 Performance test of sorptive building materials of reducing indoor air pollution with small chamber – Measurement of adsorption fl ux with supplying constant concentration of contaminant air of VOC and aldehydes without formaldehyde


JIS A 1911 2006 Determination of the emission of formaldehyde for building materials and building related products – Large chamber method

JIS A 1912 2008 Determination of the emission of volatile organic compounds and aldehydes without formaldehyde for building materials and building related products – Large chamber method

JIS C 9913 2008 Measuring method of volatile organic compounds and carbonyl compounds emissions for electronic equipment – Chamber method

ECMA - 328

Table 6.1 Continued



JIS X 6936 2005 Information technology – Offi ce equipment – Measurement of ozone, volatile organic compounds and dust emissions rate from copiers, printers and multifunction devices

ECMA - 328

Test protocols including methodology and pass/fail criteria

DIBt – German Institute of Building technology

AgBB – Comm. for Health Related Evaluation of Building Products

Health - related evaluation procedure for VOC and SVOC emissions from building products

German contribution to European discussion on the CPD. Basis for prEN 15052

California Dept of Health Services

The ‘ Collaborative for High Performance Schools (CHPS) Section 01350

Standard practice for the testing of volatile organic emissions from various sources using small - scale environmental chambers

CEN prEN 15052 Resilient, textile and laminate fl oor coverings – Evaluation and requirements of VOC emissions

‘ Emissions screening ’ or ‘ content ’ test methods

CEN EN 120 Wood - based panels. Determination of the formaldehyde content. Extraction method called the ‘ Perforator method ’ .

CEN TC351 Working Draft (Annex B)

Construction products - Assessment of emissions of regulated dangerous substances from construction products - Determination of emissions into indoor air

GUT - Gemeinschaft Umweltfreundlicher Teppichboden

GUT test method for the screening of VOC - emissions from textile fl oorcoverings

VDI - Verein Deutsche Ingenieure

2083 - 17 Draft Cleanroom technology - Compatibility with required clean lines class and surface clean lines

Table 6.1 Continued



ASTM WK proposed Standard Practice for: Micro - Scale Test Chambers for Screening Vapor - Phase Organic Emissions from Materials/ Products

VDA – Verband der Automobileindustrie

Method 278 Thermal desorption analysis of organic emissions from car trim components

German car industry method

CEN ISO 11890 - 2 Paints and varnishes – Determination of volatile organic compound (VOC) content. Part 2 Gas Chromatographic method.

US EPA Method 311 Analysis of hazardous air pollutant compounds in paints and coatings by direct injection into a GC

Table 6.1 Continued

6.2.3 Method Variability or Uncertainty

For emissions testing to be accepted as a meaningful and necessary part of product quality assessment, relevant test methods must ensure acceptable uncer-tainty. Any associated products standards, incorporating pass/fail criteria, must also take into account the actual uncertainty of the standard methods specifi ed. A relatively detailed summary of some of the major potential causes of error in the multistep materials emissions testing process is presented later in this chapter (see Section 6.6.2 ). For an emissions test standard/protocol to be robust and useful, it must take into account all of these issues and include suffi cient guid-ance to ensure that a competent laboratory can achieve results within the expected uncertainty limits.

6.2.4 Nonuniformity of Test Methods

The challenge of harmonizing multiple existing and nonuniform emissions test methods/protocols is perhaps the most diffi cult of all to address because of the vested interests that are invariably involved. However, if things are left unchecked,

6.3 Regulations, Standard Methods and Test/Certifi cation Protocols 131

manufacturers will continue to need product certifi cation according to several dif-ferent emissions test protocols if they want unrestricted access to multiple markets. Ultimately, the excessive cost burden of multiple repeat tests could force smaller producers out of business which would reduce consumer choice and be counter - productive.

Various (brave!) initiatives are currently underway to try to address this issue.

• The International Organization for Standardization ( ISO ) Technical Committee ( TC )146 sub - committee ( SC )6 maintains the main EN ISO 16000 - series of test methods for emissions testing and associated IAQ (see Table 6.1 ) and is working in conjunction with ISO TC 22 (cars) to harmonize associated test methods for in - vehicle air quality monitoring and emissions from car trim components.

• Following on from ECA - IAQ Rpt. #24 (2005) the EC expert group ‘ European Collaborative Action on Indoor Air and its Impact on Man ’ is spearheading a project to harmonize European test schemes for construction products. The four major labeling schemes participating in this work are: the German AgBB scheme; the Finnish M1 protocol; the Danish Indoor Climate label (Fox, 2007 ); and the French AFSSET (2004) scheme.

• The European Standards Agency (CEN), under TC351, is also working on harmonized horizontal test protocols which could be used to fulfi l ER3 of the construction products directive under EC mandate M/366.

6.3 Regulations, Standard Methods and Test/Certifi cation Protocols

Before reviewing the range of standard methods and protocols that are already published or that are in the later stages of development, it is important to under-stand the fundamental differences and relationship between legislation or regula-tions, analytical test methods and certifi cation protocols and product standards:

Legislation is passed by state, national or international regulatory agencies and creates a requirement to test emissions from one or more types of product/mate-rial. (Industry specifi c voluntary labeling schemes can create a similar requirement for emissions testing.) Examples of relevant regulations include:

• The 1989 EC Construction Products Directive (CPD) – Essential Requirement (ER) 3 of the CPD states that ‘ The construction work must be designed and built in such a way that it will not be a threat to the hygiene or health of the occupants or neighbors, in particular as a result of any of the following: � the giving off of toxic gas; � the presence of dangerous particles or gases in the air; � the emission of dangerous radiation; � pollution or poisoning of the water or soil;


� faulty elimination of waste water, smoke, solid or liquid wastes; � the presence of damp in parts of the works or on surfaces within the works. ’

• The Californian ‘ Collaborative for High Performance Schools ’ requires contractors bidding for the construction of new schools and other public building projects to guarantee that they will only use materials whose emissions have been tested and certifi ed by an accredited laboratory using the 01350 test protocol (see Table 6.1 ).

• German national guidelines require fl ooring products to have their emissions tested/certifi ed by an accredited laboratory using the AgBB protocol (see Table 6.1 ), before they can be sold in Germany.

• REACH: REACH was adopted by the European Parliament at the end of 2006 and came into force on 1st June 2007. It implements a uniform legal system for all chemicals inside the European Union and European Economic Area and covers all phases of product life: manufacture, use and disposal. REACH will impact an estimated 80% of manufacturing industry and requires all chemicals that are used in quantities above 1 tonne per year to be registered. Chemical manufacturers, importers and downstream users will need to prepare technical dossiers and/or ‘ Chemical Safety Reports ’ depending on tonnages. If the chemical substance is classifi ed as ‘ dangerous ’ or a ‘ substance of very high concern ’ ( SVHC ) the Chemical Safety report must include exposure scenarios. Moreover, if articles or preparations contain SVHC chemicals at levels above 0.1%, which could be released under normal conditions of use (intentional or unintentional emissions) and which could therefore present a risk to human health or the environment, this must be covered in one or more of the exposure scenarios. It is anticipated that this requirement will lead to a signifi cant expansion of chemical emission testing from consumer goods generally. Guidance on REACH requirements for substances in articles and preparations and associated emission test methods has been published by the European Chemical Agency (ECHA, 2008a ). Many of the methods listed were originally developed for construction products.

It is interesting that regulations such as these are not effective in isolation. For example ER 3 of the CPD, requiring that chemical emissions from products used indoors must not adversely impact the indoor environment, became European law in 1989. However, it has not yet been effectively implemented. This is partly because there has not, hitherto, been a validated and broadly applicable (horizon-tal) standard test method available to carry out emissions measurement in com-pliance with the regulation.

Standard Test Methods on the other hand, specify procedures for carrying out measurements. In the case of emissions testing, methods are usually broken down into multiple separate sections covering sampling frequency, sample collection and preparation, emissions testing and vapor analysis (see below.) Test methods may also include lists of target compounds or compounds that are commonly

6.4 Emissions Test Methods for VOCs: An Overview of Basic Principles 133

found in emissions from conventional building products, but this is not always the case.

Continuing with the above example of ER3 of the European Construction Prod-ucts Directive: horizontal emissions test methods (i.e., applicable to multiple product types) have been developed by CEN and ISO for compliance with the CPD and are now available as parts 6, 9, 10 and 11 of EN/ISO standard 16000 (see Table 6.1 ). Furthermore, a program of work led by CEN TC351 has now been instigated under European Council Mandate M/366 to amalgamate and validate these stan-dards. Once this work is completed (estimated: 2010) it is understood that chemi-cal emissions testing will become a mandatory part of CE marking for construction products.

Test or Certifi cation Protocols and/or Product Standards are similar to standard test methods but include additional criteria (requirements). Typical test protocols or product standards specify which emission test method should be used, when it should be carried out (for example; 3 days, 12 days or 28 days after the sample has been placed in the test apparatus), a list of target compounds (i.e., which com-pounds or groups of compounds are to be measured) and pass/fail criteria (i.e., limit levels for emissions above which a product would fail the certifi cation process because one or more compounds are being emitted from that product material at an unacceptably high rate.)

6.4 Emissions Test Methods for VOC s : An Overview of Basic Principles

6.4.1 Standard test Methods for Formal Evaluation and Certifi cation of Emissions

As outlined above, conventional emission test methods for vapor - phase organic chemicals attempt to measure emissions under simulated real - world con-ditions. These methods are typically broken down into several sections and may include:

• Statistical analysis of when and where to collect the samples for emissions testing ( – to make sure that the measurements will be representative).

• A detailed guide to sample collection and the preparation of test specimens from the material or product samples collected, including sample storage, transport, cutting, application, curing, etc. Note that the preparation procedure will vary greatly depending on the type of product involved and how it is used in building construction/decoration.

• Specifi cation of the type of test chamber or cell required for emissions measurements together with associated performance criteria. Guidance is usually also given for the positioning of test specimens within the chamber


or cell and the conditions under which measurements should be made (see Table 6.2 ). Note that, once installed in a building, most construction products only have one surface exposed. Therefore, in order to evaluate a product with respect to its expected contribution to indoor air pollution after installation, product edges and rear surfaces are usually masked during emissions testing.

• Details of the type of samplers that should be used for collecting the emitted vapors from the exhaust stream of the test chamber or cell and the recommended times for testing.

• The method of analyzing the vapor samples collected from the test chamber/cell exhaust stream plus guidance as to how to use that data to derive an area - or volume - specifi c emission rate for the product under test. Analytical methods used for testing vapor - phase emissions typically fall into two main categories:

� Use of in - sampler derivatization of formaldehyde and other carbonyls using DNPH cartridges followed by HPLC analysis of the derivatives.

� Use of sorbent tubes to trap a broad range of organic vapors ranging in volatil-ity, for example, from n - hexane ( n - C 6 ) to n - hexadecane ( n - C 16 ) and above, followed by analysis with TD – capillary GC and MS detection, with or without the additional use of fl ame ionization detectors (GC – FID).

Data are typically reported as mass emitted per unit surface area (or volume) per unit time. This is known as the area (or volume) specifi c emission rate ( μ g/m 2 /h.) Alternatively, some test protocols require results to be presented in terms of vapor concentration, either in the chamber/cell itself or in a specifi ed model (reference) room. NOTE that specifi c emission rate data can be converted to concentration data (and vice versa) by means of calculation.

This type of emissions test method is widely required for formal product certi-fi cation because it generates analytical data that relates most closely to real - world use. However, even from the above brief description it is clear that formal emis-sions testing methods are lengthy and relatively complex, which translates to rela-tively ‘ expensive ’ for manufacturers. The time required (several days, even weeks) also precludes the use of these standard procedures for quality control of produc-tion, quick product screening or for convenient testing of prototype materials or products under development. Thus, while there are few that would contest the use of conventional emissions testing (as outlined above) for formal product certifi ca-tion, there is an additional need for complementary more - rapid and less - expensive emissions screening methodology that could be used by manufacturers and others for routine quality control and product development.

6.4.2 Secondary or ‘ Screening ’ Methods for Materials Emissions

Recently, the need for additional secondary ( ‘ initial ’ or ‘ screening ’ ) methods, to complement formal (certifi cation) emissions test methods, has been recognized

6.4 Emissions Test Methods for VOCs: An Overview of Basic Principles 135

Table 6.2 Conditions for material emissions testing using chambers or cells.

Parameter Range Comments

Temperature 23 ° C (Europe/US), 28 ° C (Japan) Correlates to typical indoor conditions

Humidity 50% R.H.

Test time 3 days and 28 days (CEN and ISO), 10 days and 12 days (US protocols). 24 - hour tests proposed in some cases

Allowed size Small chambers typically range in volume from 20 l to 5 m 3 . Cells are much smaller – typically 35 cm 3

Larger chambers allow larger (more representative) samples to be tested, but need much longer to equilibrate.

Material of chamber/cell construction

Nonemitting, nonsorptive surfaces – typically polished stainless steel or glass

Air supply (l/min) 0.2 to 1.0 (cells), 0.5 to 20 (chambers)

Equilibration time Standard methods typically specify a minimum of 24 hours for cells and 72 - hours for chambers.

In practice, equilibration times can be as low as 15 – 20 minutes for cells.

Surface area Typically < 200 cm 2 in cells, but up to 10 000 cm 2 in larger chambers

Multiple tests may be required for inhomogeneous products such as natural wood if using smaller chambers or cells

Air distribution Air velocity is never completely uniform across the sample surface

Though surface air velocity does vary in emission cells as in chambers, cells do preclude the variability of sample orientation

Recovery > 95% toluene and n - C 12 (cells) and > 80% toluene and n - C 12 (chambers)

Used as a check against ‘ sink effects ’ within the chamber/cell

Target analyte range Formaldehyde and VOCs ( n - hexane to n - hexadecane)

Very volatile and semi - volatile organic compounds also considered

Samplers for vapors in cell/chamber exhaust

DNPH cartridges for formaldehyde Tenax TA or alternative sorbent for VOCs

When using Tenax TA tubes for VOCs, max volume restricted to 5L

Detection limits 2 μ g/m 3 vapor concentration in exhaust stream from chamber/cell

Analysis technique Formaldehyde – HPLC of DNPH derivative VOCs – TD - GC - MS/FID


by various regulators and standards organizations. Screening methods have been proposed, for example, by:

• ECHA for release of chemicals from articles under REACH (ECHA, 2008a ); • EC mandate M/366 for the CPD; • ASTM (ASTM, 2008 ); • various industries (GUT, 2008 ; VDI, 2008 ).

Under M/366 and the CPD, one suggestion is that screening or initial test methods could be used to demonstrate that a product or material can be classed as ‘ without further testing ’ ( WFT ), allowing manufacturers to short - cut the full certifi cation process.

Regulators on both sides of the Atlantic are also keen to provide manufacturers with a cost - effective means of demonstrating that a product continues to conform to emissions limits in between formal certifi cation tests. From industry ’ s perspec-tive; screening methods would also facilitate in - house checks on batch - to - batch product uniformity/conformity, at - line quality control, confi rmation of the results of external certifi cation tests and demonstration of emission profi le consistency across a product range (e.g., demonstrating identical emission profi les from dif-ferent colored versions of the same product, to minimize the requirement for formal certifi cation testing). Practical and quick screening methods would also contribute to the development of low emissions products by allowing manufactur-ers to test new products during development and by enabling them to compare their own materials against recognized ‘ best - in - class ’ products.

At the current time, few analytical methods have been developed for emissions screening from construction products. Examples of the types of method available include EN ISO 11890 - 2 for measuring the organic content of paint, VDA 278 for screening car - interior trim components for volatile and semi - volatile (fogging compound) emissions, the GUT screening method for fl ooring emissions and VDI 2083 - 17 for screening emissions from products used to construct and furnish clean - room fabrication facilities (see Table 6.1 ). Screening methods used to date for VOCs have typically involved either direct GC analysis of the ‘ volatile ’ content of products which are applied as liquids (paints, coatings, etc.) or direct thermal desorption/extraction of small samples of solid or liquid - applied products and materials.

The challenge for developing useful screening methods is to fi nd something that meets the criteria of speed, simplicity and cost, but that still produces data which correlates with results from standard emissions testing methods. The issue with organic content testing is that results rarely correlate to emissions from the respective products after they are applied/installed. Manufacturers of paint or paint additives, for example, can design products that contain solvent, but in which that solvent is encapsulated such that it can never escape to the indoor environ-ment. Content testing is also unsuitable as a guide to emissions from most com-posite products. Signifi cant discrepancies may additionally be observed between data from direct thermal desorption and standard emissions testing methods. In this case, the differences are primarily the result of:

6.5 The Total-VOC Debate 137

• The high temperatures used for direct thermal desorption/extraction. Note that emissions testing is conventionally carried out at room temperature (Table 6.2 ) whereas VDA Method 278 for example requires desorption temperatures of 90 and 120 ° C for volatiles and ‘ fogging ’ compounds respectively.

• Direct thermal desorption is usually carried out on bulk samples, that is, with emissions from multiple surfaces rather than just one exposed surface.

• Direct thermal desorption is normally limited to relatively small sample sizes ( < 1 g) which may not be representative of the product/material as a whole.

Low temperature/surface - only ‘ thermal ’ extraction/microchamber methods have recently been evaluated to see if the results obtained provide better correlation with standard emissions tests (GUT, 2008 ; VDI, 2008 ; Schripp et al ., 2007 ; HEMICPD, 2009 ) and whether or not they can be used to accurately and rapidly predict product performance with respect to chemical emissions after 3 or 28 days. Some of these studies are still on going but results have been very positive and have already led to the consideration/development of several low temperature, surface extraction methods for emissions screening (ASTM, 2008 ; GUT, 2008 ; VDI, 2008 ).

6.5 The Total - VOC Debate

Many standard emissions test methods and protocols require ‘ measurement ’ of the TVOC emission rate in addition to that of individual compounds. This is usually obtained from the GC data by summing the masses of every individual analyte which elutes within a particular range (typically n - C 6 to n - C 16 ), on a non-polar capillary GC column. Detailed procedures for this vary, with some protocols calling for individual measurements to be made by FID, others by MS. Similarly some protocols require individual compounds to be calibrated using ‘ authentic standards ’ (i.e., standards of each specifi c compound found) while others allow measurements of all the individual compounds which contribute to a TVOC data point to all be calibrated as toluene (that is, in ‘ toluene equivalents ’ ).

Attitudes to TVOC measurements vary widely. Some like the concept of a ‘ TVOC ’ parameter because it is simple to understand as a pass/fail criteria. However, it is not a reliable indicator of product safety or acceptability unless it is specifi cally applied to comparisons of very similar products, for example, different batches of the same material or products within a range that are essentially the same, but may have a different color or patterned fi nish.

The main limitation with TVOC as a measurement is that the toxicity of organic chemicals varies over six orders of magnitude. Some organic chemicals, common ethanol/alcohol for example, have very low toxicity; an adult human could swim in ethanol for a short while and survive unscathed. At the other extreme however, there are compounds which are fatal if inhaled at low ppb concentrations. Thus a


single TVOC reading does not, in itself, give a reliable indication of the potential health risks of product emissions; in other words; a low TVOC result does not necessarily mean that a product is safe because one or more of the individual components within that low TVOC measurement could be a dangerous carcinogen or reprotoxin.

Moreover, variation in the methods of measuring TVOC, FID vs MS detection, toluene vs. authentic calibration standards, summing individual peaks or measur-ing the entire chromatographic area, etc., plus the choice of chromatographic integration parameters, can have a dramatic effect on the TVOC result. Therefore, even when using TVOC as a tool for comparing similar products, care must be taken to ensure that all the TVOC data have been collected in exactly the same way.

6.6 Standard Methods and Protocols for Emissions Testing: Current Status

A list of the most prominent materials emissions standards and test protocols is presented in Table 6.1 . The current level of method development activity is such that it is diffi cult to be confi dent that all relevant standard methods, both published and in later stages of development, have been included, but care has been taken to try to make this list as comprehensive as possible.

A key to the acronyms used in Table 6.1 is as follows:

CEN Committee European de Normalization – European Standards Agency ISO International Standards Organization ASTM American Society for Testing and Materials JIS Japanese Industrial Standards EN European Norm prEN Provisional European Norm US EPA United States Environmental Protection Agency WK ASTM notation for working draft

6.6.1 Typical Conditions for Emissions Testing Using Chambers/Cells

Conventional materials emissions test methods are usually carried out using small test chambers or cells with samples prepared and presented to the air fl ow within the chamber/cell as they would be to room air in real use. Operating parameters (temperature, humidity, etc.) are selected to simulate the indoor environment and are summarized in Table 6.2 .

The fundamental difference between test chambers and cells is that samples are placed within test chambers whereas test cells are placed onto the emitting surface of a product or material. Emissions cells are limited to planar materials or applied paints/coatings and cannot be used for moulded or nonplanar materials. However, they are inherently quicker and easier to use than chambers and eliminate the

6.6 Standard Methods and Protocols for Emissions Testing: Current Status 139

variable of sample orientation. A detailed comparison of emissions test chambers and cells has been reported previously (Wolkoff et al ., 2005 ). NOTE: Micro - cham-ber - type devices used for low - temperature surface emissions screening can often be used as either mini chambers or cells.

6.6.2 Standard Methods: What Can Go Wrong?

Although regulators and consumers are pushing for increased materials emis-sions testing from a public health and environmental protection perspective, relevant standard methods must be capable of delivering consistent and reliable data before industry will be happy to accept it as a necessary and useful part of routine product quality control. For the reasons given above, formal materials emissions testing is a relatively complex, multistep process. Understandably manufacturers want to be confi dent that accredited laboratories will produce con-sistent results for the same sample (within reasonable uncertainty limits) before willingly accepting the incremental costs emissions testing will incur. Given that emissions testing is a multistep process and that the methods can be applied to such a diverse range of materials, there are many areas that are open to inter-pretation and where intra - and inter - laboratory discrepancies can creep in. A detailed analysis of uncertainty in emissions testing has recently been reported (Howard - Reed et al ., 2008 ) and a summary of the key issues contributing to vari-ability is presented below.

6.6.2.1 Effect of the Emission Mechanism One of the most signifi cant potential causes of uncertainty in emissions testing occurs if emissions are required to be measured from liquid - applied products during the drying/curing stage. Materials emissions are typically the result of either evaporation (external diffusion) or internal diffusion/migration of volatiles within the bulk material. Evaporative emissions dominate during the drying/curing of liquid - applied products and are signifi cantly affected by air velocity/fl ow at the surface of the sample. Other parameters such as temperature, time and humidity similarly have a very signifi cant impact on evaporative emissions. The extreme susceptibility of evaporative emissions data to so many variables, includ-ing nonuniformity of air velocity in small - chambers and cells, makes it almost impossible to get good inter - laboratory reproducibility and makes the value of any such data highly questionable.

Emissions measurements are much more reproducible and meaningful if the dominant mechanism is internal diffusion, for example, for solid materials or for liquid - applied products after the drying/curing stage. Although storage/transport/preparation conditions and all analytical parameters must still be rigorously con-trolled (particularly temperature), emissions controlled by internal diffusion are much less susceptible to variation in time or surface air velocity than evaporative emissions. Several successful round - robin, multi - laboratory studies and compari-sons of various emission test chambers/cells have shown satisfactory agreement


(Jann et al ., 1999 ; Hansen et al ., 2000 ; Afshari et al ., 2003 ; xAkutsu et al ., 2000 ; Roache et al ., 1996 ; Daumling et al ., 2008 ).

6.6.2.2 Collection and Transport of Samples Plus Homogeneity Issues Most standard procedures recommend test samples be collected/shipped directly from the manufacturer or retail outlet in the normal product container with addi-tional air tight, nonemitting packaging if required. From the point of sample selection, it is critical that emissions are minimized/eliminated by using air - tight packaging and that no potential contaminants are allowed to ingress to the sample from the outside world. Contamination is a particular issue during transportation where high concentrations of volatile fuel components may be present.

Products most likely to be subject to signifi cant inhomogeneity include natural materials such as timber (treated or untreated) and some painted or printed sur-faces. Knots in wood and particularly dense areas of color can cause relatively intense, localized sources of emissions. Potential uncertainty due to inhomogene-ity is best addressed by using larger samples or by carrying out multiple emission tests on smaller pieces of the same sample.

6.6.2.3 Potential Variables Associated with Testing Materials Using Emissions Chambers/Cells: Edge Effects, Sample Orientation and Sample Storage Between Tests Preparation of test specimens from solid samples invariably requires a portion of the sample to be cut out, for example, from the middle of the original roll or sheet of material. Freshly cut edges are notoriously high emissions sources and failure to adequately seal these and exclude them from the test is a major potential source of error. (Another advantage of emission test cells versus conventional chambers is that they automatically exclude any potential edge effects in most cases.) Simi-larly, surface emissions testing requires the rear surface of the material (the surface that will not be exposed once the product is installed in a building) to be sealed and excluded from the test. This is always a consideration for emission test chambers but can be an issue for both cells and chambers in the case of porous or permeable materials.

Emission rates may also be affected by sample positioning and orientation within a test chamber and the relation to the air ventilation and mixing system. This is a particularly critical issue for liquid applied samples tested during the drying/curing phase where evaporative emissions dominate (see above). However, care must always be taken to ensure that samples are positioned in the chamber so as not to impede the air fl ow and to ensure that vapors emitted by the sample are adequately mixed with the chamber air before reaching the exhaust outlet. Emission cells applied to planar surfaces are not, generally speaking, affected by these issues.

Storage of samples between tests is another potential source of error. In an ideal world, samples should be kept in the chamber or under the cell throughout an emission test. However, given that many standard methods and protocols require a 2 - or 4 - week evaluation and that each fully confi gured emissions test

6.6 Standard Methods and Protocols for Emissions Testing: Current Status 141

station – chamber/cell plus ancillary support equipment (pure air supply, humidi-fi cation, temperature control, etc.) – costs from ∼ US$10 000 to > $100 000 ( ∼ € 8500 to > € 85 000), this is, in most cases, prohibitively expensive. Samples are thus often removed from the test chamber in between tests (for example between tests at 3 and 28 days) and only replaced in the chamber or cell 72 - or 24 - hours respectively before a measurement needs to be made. Control of the sample storage conditions during this period is critical. Temperature, humidity and levels of atmospheric contaminants in the storage containers must be maintained at, or as near as pos-sible to, the conditions specifi ed for the emission test itself. Storage containers must also be well ventilated and care must be taken to prevent cross - sample/product contamination.

6.6.2.4 Sink Effects Emission test chambers/cells and associated air supply and ventilation/air mixing equipment must be nonemitting/ - absorbing and designed to minimize still air volumes and risk of ‘ sink effects ’ within the chamber/cell, that is; the apparatus must be designed to prevent condensation of organic vapors on the inner surfaces of the chamber or cell. Tests for higher boiling organic emissions are particularly prone to sink effects which cause emission rates to be underestimated and may also increase risk of background contamination in subsequent tests of other samples with the same apparatus. Most well recognized emissions test methods/protocols require assessment of sink effects in the apparatus by introduction of pure standards of toluene, n - C 12 and key compounds of interest (especially if they include polar compounds or analytes less volatile than n - C 12 ). Such checks are an essential part of analytical quality assurance for material emissions testing methods.

6.6.2.5 Target Analytes and System Calibration Extensive lists of target analytes, if associated with stringent calibration require-ments (e.g., requiring authentic standards for every identifi ed compound) can sometimes increase test costs without necessarily reducing analytical uncertainty. In fact, diffi culties in sourcing and preparing large numbers ( > 50) of authentic compounds in mixed standards and the time taken for subsequent multilevel calibration of the analytical system with all these standards, can even add to measurement uncertainty in extreme cases. For example, if the time taken for multilevel calibration extends over several days, it is possible for the system response to have drifted signifi cantly over the period reducing confi dence in the subsequent data. Other diffi cult to characterize contributions to variability can also creep in, for example, analyte interactions and the general stability of mixed standards.

While there is an understandable demand from regulators and consumers to quantify as many emitted compounds as possible, as accurately as possible, there needs to be a balance between the desire to use authentic standards in each case and the need to ensure that the calibration process is practicable to carry out over a reasonable time period. Calibrating with fewer compounds, for example a mix


containing ∼ 10 compounds, including components which are broadly representa-tive of the range of chemical compounds of interest, can be achieved more repro-ducibly, quickly (and affordably) and thus lead to a higher degree of confi dence in subsequent analytical data.

Care must certainly be taken to ensure that any and all compounds, specifi ed as target analytes in emissions test standards/protocols, are compatible with the analytical methodology used. For example; many standard emission test methods have evolved to focus on GC - compatible VOC components, ranging in volatility from n - hexane to n - hexadecane ( n - C 6 to n - C 16 ). The rationale for this is that more volatile or reactive emissions should disappear so quickly after product installation that they are unlikely to affect the health or quality of life of building occupants. However, this takes no account of multilayer products where VVOC emissions may be slowed by a coating, or the potential for very volatile organics to continue to be produced by chemical processes within a material (e.g., degradation of an additive) long after installation.

Similarly, SVOC emissions are traditionally considered to present a lower risk because their vapor - phase concentration in ambient air is, by defi nition, very low.

These considerations have led to most standard emissions test methods and protocols focusing on Tenax TA sorbent for trapping/sampling vapor - phase organ-ics from the exhaust stream of chambers/cells. Tenax is a relatively weak sorbent, offering only 6 l retention volume for n - hexane in a standard tube at 20 ° C. However, it has other useful properties such as low artifact levels, effi cient desorption/release characteristics and excellent hydrophobicity (water passes through almost unretained provided the sorbent trap/tube is not cooler than the sampled air/gas). Provided the volume of vapor pumped onto the sorbent tube is kept below 5 l, most GC - compatible organics within the n - C 6 to n - C 16 volatility range will be quantita-tively retained by the Tenax during vapor sampling and will be effi ciently desorbed during subsequent analysis. However, restricting the vapor trapping/sampling medium to Tenax means that more volatile compounds – n - pentane, vinylchloride, propanol, methyl chloride, etc. – cannot be included among lists of target analytes. It also means that vapor sampling volumes must be restricted to 5 l or less. More information on the potential toxicity of some common semi - volatile additives such as phthalates (Larson et al ., 2008 ; ECHA, 2008b ), has also caused a recent increase in demand for the accurate measurement of much higher boiling compounds, species that cannot be completely recovered from Tenax TA.

These developments have led to increased focus on additional sorbents, which can be used in conjunction with Tenax TA (i.e., in the same sorbent tube) in order to increase the target analyte volatility range without increasing measurement costs. (ISO DIS 16000 - 6, 2008 ; CEN TC 351 WG 2, 2008 )

6.6.2.6 Chromatographic Integration and Summation Limit Levels Potential concerns with TVOC (and total SVOC [TSVOC]) have been discussed above. For TVOC or TSVOC data to be of any value, even for the intercomparison of different batches of the same product, care must be taken to ensure that exactly the same procedure is used for each TVOC or TSVOC measurement.

6.8 Concluding Remarks 143

Other potential causes of discrepancy include variable integration results (Oppl, 2008 ) resulting from components which are not adequately chromatographically resolved. This can be particularly signifi cant when trace target compounds are required to be measured at levels that are close to system or method detection limits. New enhanced software data - processing tools are becoming available to address this (Rosser et al ., 2008 ), but if such data is interpreted by unskilled ana-lysts without access to appropriate advanced data processing tools, the potential for error remains.

6.7 Confi dence Limits for Emissions Test Data for Individual VOC s

As has been described, conventional material emissions testing is a multistep process involving sample selection, transport and storage, emissions testing using chambers/cells, vapor sampling and, fi nally, TD – GC – MS – FID analysis. Each stage of this process introduces some level of uncertainty. By way of comparison it is worth noting that even the most rigorous air monitoring methods for individual vapor - phase VOCs, for example EN 14662 - 1 (2005) for ambient benzene, quote confi dence limits in the order of 15%. This methodology involves pumped sam-pling of ppb - level vapor onto sorbent tubes with subsequent TD – GC( – MS) analysis and is analogous to the latter stages of material emissions testing except that only one compound is involved. A reasonable estimate of the confi dence that could be expected for pumped tube sampling and TD – GC – MS – FID analysis of multiple ppb - level VOCs (for example using EN/ISO 16017 - 1 or ASTM D 6196) is 25 – 30%.

If the additional variables of sample selection, transport and preparation and chamber/cell emissions testing are taken into account, overall uncertainty for emissions testing methods is likely to be in the range 30 – 50%, even with best practice. Pass/fail criteria in associated test protocols and product standards, must take this into account.


Recent regulatory developments at state, national and international level suggest that certifi cation of products, according to their emissions, is expected to become mandatory across much of the developed world for building materials and related products. This will affect both manufacturers and importers.

Harmonization, simplifi cation and validation of relevant standard test methods/protocols are all essential steps if emissions testing is to become as widespread as regulators and consumers desire, without burdening industry with unjustifi ed extra cost. Failure to achieve this could reduce confi dence in materials emissions testing generally and thus undermine the overall objective. It could also jeopardize

144 6 Testing Emissions of Organic Vapors from Building Products to Indoor Air

the survival of smaller specialist producers and those from the developing world as well as having an adverse impact on product diversity, none of which is socially desirable.

Parallel development of broadly - applicable and meaningful emissions screening methods will be another important factor in facilitating the development of improved ‘ low - emission ’ materials and in allowing emissions testing to become an accepted part of routine in - house quality control during product manufacture. Similar attention needs to be paid to broadening the base of competent and accred-ited emissions test laboratories to ensure competitive rates for certifi cation testing.

Perhaps the greatest challenge will be to achieve consensus on harmonization of standard methods/protocols. National standard methods, or those adopted by one specifi c industry group or labeling scheme, are often defended vigorously by the proponents of each individual scheme, unwilling to modify a procedure in which they, and their user - base, will have made signifi cant investment. However, failure to address this issue head - on and as soon as possible could lead to the promulgation of more and more independent emission test methods/schemes which will ultimately increase the burden on manufacturers wanting to supply into more than one geographical or market sector. Current international efforts in this area, for example the program of work on horizontal emissions standards for construction products planned by CEN TC 351 in response to EC Mandate M/366, should be warmly welcomed and further liaison with other geographies and industry groups encouraged.

Material emissions testing is also beginning to move out from the construction products sector and into furnishings, furniture and other equipment. Given the recent adoption of REACH in Europe and the Green Chemical act in California and the associated focus on chemical emissions from consumer goods in general, this trend is set to continue. If the construction products industry and associated test laboratories have already ‘ blazed a trail ’ in terms of developing reliable and cost - effective test methods, it is likely that these methods will be extrapolated, where appropriate, to products and materials in other sectors.

Acknowledgments

I would like to acknowledge the contribution of Dr Derrick Crump, BRE, UK and Prof Tanabe, Waseda University, Tokyo, Japan in preparation of this text.

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Standard Test Methods for the Determination of VOC s and SVOC s in Automobile Interiors Michael Wensing

7

7.1 Introduction

Inside air in automobiles is infl uenced by a large number of different factors. Air pollutants from various sources inside and outside the automobile can effect the inside air composition, see Figure 7.1 . It is important to take into consideration the possible exposure of vehicle occupants since automobile drivers can spend a considerable part of the day in the vehicle, as results of relevant studies in the USA have shown. One study, for example, states that most people spend one to four hours per day in the automobile (Weisel, Lawrik and Lioy, 1992 ). According to another source (Park et al ., 1996 ) adults spend on average around 7% of their day (approx. 1.7 h) in a motor vehicle. One important type of air pollutant is organo-chemical compounds. Their occurrence is linked to the interior components, such as seats, dashboard, headliner and carpeting, which are made of a wide range of different materials. These materials vary in their composition as well as in their internal chemical structure. To obtain their necessary and desirable characteristics and to facilitate their production a large variety of organochemical substances needs to be used, such as plasticizers, fl ame retardants and other additives (Wensing, Uhde and Salthammer, 2005 ). After completion of the production process, these materials emit chemical compounds with a wide range of volatilities into the air inside the car: these substances may be roughly divided into volatile organic compound s ( VOC s) and semi - volatile organic compound s ( SVOC s). High temperatures encourage those emissions which can cause undesirable effects in the car ’ s interior. On the one hand these can cause car users to complain of annoy-ing odors or even to experience health problems, while on the other hand the semi - volatile part of the compounds has been observed to condense as a ‘ fogging ’ fi lm on the inner side of the windscreen. Together with soot and dust particles, this fi lm impairs the transparency of the windscreen. The driver ’ s view is further impaired if the screen has a small inclination (Eisele, 1987 ; M ö hler and Sch ö nherr, 1992 ; Munz et al ., 1994 ). To reduce these disadvantageous effects and to eliminate them as far as possible, comprehensive and reliable information about the types


147

148 7 Standard Test Methods for the Determination of VOCs and SVOCs in Automobile Interiors

of organic compounds in the inside air of automobiles and also their concentra-tions, time behavior and origin is required.

In the meantime a large number of studies concerning the measurement of organochemical compounds inside the automobile have been published (e.g., Fine, Reisch and Rounbehler, 1980 ; Smith and Baines, 1982 ; Zweidinger et al ., 1982 ; Dropkin, 1985 ; Ullrich, Seifert and Nagel, 1992 ; Fromme et al ., 1998a, 1998b ; Fedoruk and Kerger, 2003 ; Brown and Cheng, 2003 ; Zhang et al ., 2006 ; Yoshida and Matsunage, 2006 ; Yoshida et al ., 2006a, 2006b ; Meininghaus et al ., 2007 ; Chien, 2007 ). These measurements were carried out under various conditions: outdoors or indoors, as well as in motion. In some cases the temperature inside the car was increased using a heater. Air samples were simply taken directly from the car ’ s interior or through a covered - up window - opening by means of an adsorption tube and a small pump. The samples were analyzed by GC with various detectors. The procedure for examining VOCs followed a similar procedure to that for the semi - volatile ‘ fogging ’ substances (SVOCs). The vehicles were either in use or parked outdoors with their windows closed and exposed to the sun. As a rule the fogging fi lm which developed on the inside of the screen was scraped off with a clean razor blade and then examined by GC and IR spectroscopy (Carter, Jensen and McCal-lum, 1987 ; Nranian, McCallum and Kelly, 1987 ; Munz et al ., 1994 ). On the basis of the results from such studies many car manufacturers are today embarking on a strategy of avoiding unwanted emissions right from the very start of the develop-ment and production of new automobiles. many test methods exist for the selection of low - emitting materials and interior components (VDA, 1992, 1994, 1995, 2002, 2005 ; Meyer et al ., 1994 ; Toyota, 2003 ; Hoshino et al ., 2005 ; Schripp et al ., 2007 ).

To measure VOCs and SVOCs inside complete cars under standardized condi-tions at T Ü V NORD in Hamburg a special test stand has been developed during the course of a 5 - year research and development project (Bauhof, 1994 ; T Ü V NORD, 1996 ; Bauhof et al ., 1996 ) and has also been used frequently (Wensing and

Figure 7.1 Air pollutants from different sources indoors and outdoors affect automobile inside air composition.

7.2 Conditioning of the Automobile Interior 149

Schwampe, 1989 – 2002 ; L ü ssmann - Geiger and Schmidt, 1995 ; Schmidt and L ü ss-mann - Geiger, 1996 ). The test stand investigations described below are very useful to automobile manufacturers within the framework of development projects for new automobiles or as part of the production quality assurance: They can then obtain valid data regarding possible emission sources in the vehicle interior and their emission potential. However, results from such studies on a test stand will not be automatically suitable for exposure evaluations in every case. For health - related aspects only those measured values should be used which have been obtained under conditions similar to the actual use of an automobile under a realistic exposure scenario. For the interpretation of test stand results it is very important to be aware that the air temperature and the air exchange rate in the vehicle interior can be fundamentally different in test stand investigations and during the driving situation. Depending on the ventilation system setting, the air exchange rate in the traveling vehicle is up to approx. 240 per hour (Schmidt and L ü ssmann - Geiger, 1996 ) and is therefore higher than the air exchange of a station-ary vehicle by a multiple (approx. 0.5 – 1 per hour), or that of other indoor spaces.

In other words test stand investigations can be subdivided on the basis of two fundamentally different goals; the principle focus of the test stand methods described below and the corresponding test protocol is to determine material - specifi c emission data with the aid of vehicle - specifi c emission rates ( ER ). If exposure is to be considered insofar as it affects health the boundary conditions of the investigations (air exchange, temperature) would need to be modifi ed to correspond to a realistic exposure scenario. In principle investigations of this kind are also possible once the test conditions have been correspondingly adapted.

7.2 Conditioning of the Automobile Interior

The concentration of VOCs and SVOCs in the automobile interior during the test stand procedure is essentially infl uenced by the following factors:

• age of the automobile;

• the strength of the emissions from numerous individual sources in the interior fi ttings and also from the hollow spaces with a connection to the interior of the car;

• decomposition reactions in the form of adsorption and absorption on the surface of materials, including chemical conversions (e.g., formation of salt from base and acid combinations, oxidation of aldehydes, and so on);

• gas exchanges between the passenger compartment and the atmosphere outside the car;

• contamination by external sources close to the car, from extraneous disturbing emissions from building materials and furnishings in the test bay, and from the outside atmosphere.


For a representative analysis of the emissions, these measurements should only be taken when the condition of the interior atmosphere is stable. This will be the case if certain effects that increase or decrease the concentration of emissions are balanced and are not subjected to any further change. This is why the precise conditioning of the car ’ s interior is of utmost importance in adjusting and retain-ing a defi ned state of equilibrium.

These requirements are met by standardizing the experiment ’ s set - up and pro-cedure. All vehicles are tested on a test stand and always in the same location. Only a standardized interface consisting of an aluminum sheet is used for insert-ing probes into the car ’ s interior for air sampling. The sheet is fi tted into the car ’ s open window and sealed. All connections leading into the interior are airtight. As a rule the set - up includes the following: several sensors to register the temperature at different places inside the interior and to measure the humidity, a mechanical system for air circulation, a glass probe to extract several samples of air simultane-ously (analysis of VOCs), a device to take samples of phthalates directly, and a device to enrich the fogging precipitate (SVOCs).

The test protocol logged for the experiments usually depends on the particular question which has been asked. The essential conditions for the surroundings, however, are as follows:

• The temperature of the inside air has to be adjusted to 23 ° C (ambient air tem-perature), 45 ° C and 65 ° C which are defi ned as standard temperatures for the inside air.

• The air has to be kept in constant circulation by mechanical means and thus homogenized.

• The heating of the interior is effected from outside the car by radiators.

• If large samples of air are taken from the interior of the car, ultra - pure gas must be added to balance the draw - off.

Figure 7.2 is a diagram of the test stand showing:

• provision of ultra - pure gas (A) of defi ned humidity;

• semi - automatic heating - up device with its controller (H), provision of electricity (G) and the four radiator fi elds (F) which are attached to a support above the front and the rear windows, the roof and the right - hand side window;

• a vehicle (B) to carry probes, with an arm and a standardized interface;

• various measuring instruments and data - recording equipment (C) for the con-tinuous recording of measured signals (e.g., for eleven spots for temperature measurement, three spots for humidity measurements, one FID signal and also signals from other measuring instruments);

• a system for taking air samples;

• a gate (K) for entry into and exit from the test car.

7.3 Measurement Procedure 151

Figure 7.2 Arrangement of the main components of the test stand.

Figure 7.3 shows partial views of the test stand for inside air measurements in automobiles. Detailed information about the test stand may be found in T Ü V NORD (1996) . Figure 7.4 a shows a long - term standard test cycle with measurements taken at room air temperature and at 45 ° C and 65 ° C and including artifi cial ageing periods. This cycle was used for the measurement of new auto-mobiles during a 5 - year research and development program at T Ü V NORD (1996) . For routine measurements a modifi ed short - term test protocol is in use with measurements at room temperature and at 65 ° C without ageing (see Figure 7.4 b). More than 150 different automobiles have now been tested on the basis of this short - term test protocol.

7.3 Measurement Procedure

In the interior of a modern automobile, the presence of a large variety of individual VOCs and SVOCs can be demonstrated. They can roughly be categorized as follows:

• VOCs responsible for the smell of brand new cars: alkanes, aromatic hydrocar-bons, carbonyl compounds, residual monomers, alcohols, esters, ethers, halo-genated hydrocarbons, terpenes, nitrogen and sulfur compounds;

• SVOCs responsible for the ‘ fogging ’ effect: paraffi ns, higher fatty acids and esters, phthalates, phosphoric acid esters, organosilicon compounds, haloge-nated hydrocarbons, oxygen, nitrogen and sulfur compounds.

The concentrations range from a few hundred μ g/m 3 to less than 1 μ g/m 3 of inside air. High standards of sensitivity and selectivity are required for the analyti-cal methods. The analysis is complicated by water vapor which is released from water stored in the textile surfaces when the vehicle heats up. This can have a


disturbing effect on sampling and analysis. It is therefore essential that for the analysis of the air inside cars procedures are employed which have been specially validated and also tested in practice.

7.3.1 Quantitative Determination

The standard procedures for the determination of important substance classes can be summarized:

Aromatic hydrocarbons • Sampling: charcoal tubes, type NIOSH • Sample preparation: desorption with carbon disulfi de • Analysis: GC − MS, SIM - Mode; column: DB5, 60 m × 0.25 mm × 0.25 μ m

Figure 7.3 Different partial views of the test stand for inside air measurements of automobiles.


Glycol ethers • Sampling: charcoal tubes, type NIOSH • Sample preparation: desorption with dichloromethane / methanol • Analysis: GC − MS, SIM mode; column: DB - WAX, 60 m × 0.25 mm × 0.25 μ m

Aldehydes / ketones • Sampling: derivatization with DNPH in acetonitrile • Analysis: HPLC with ODS - Hypersil/RP8e and LiChrospher 100RP - 8e • UV detection at 360 nm

Phthalic acid esters • Sampling: glass fi ber fi lter with Florisil tubes • Sample preparation: desorption with acetonitrile • Analysis: HPLC, ODS - Hypersil • UV detection at 234 nm

Figure 7.4 (a) Long - term standard test cycle. (b) Short - term standard test cycle.


Amines • Sampling: silica gel tubes • Sample preparation: derivatization with 9 - fl uorenylmethylchloroformate

( FMOC - CL ) • Analysis: HPLC, ODS Hypersil • UV detection at 263 nm

Nitrosamines • Sampling: Thermosorb N tubes • Sample preparation: desorption with dichloromethane/methanol • Analysis: GC − MS, SIM mode, PCI • Column: DB - WAX, 60 m × 0.25 mm × 0.25 μ m

The results of the quantitative measurements are given as concentrations of sub-stances, expressed in mass per unit volume (e.g., μ g/m 3 , standardized for gas under the following conditions: temperature 20 ° C, pressure 1013 hPa, dry). As a rule these measured values relate only to the time span of the sampling and the condition of the car ’ s interior at that time. If the potential emission of a substance i is to be determined, the emission rate ER must be calculated on the basis of the measured concentration:

ER g htot i= ( )Q C μ (7.1)

Q tot (m 3 /h) is the volume fl ow at which the interior atmosphere is exchanged while measuring the concentration. It is determined experimentally by tracer gas methods (e.g., T Ü V NORD, 1996 ).

7.3.2 Semi - Quantitative Determination of VOC s ( TVOC )

The VOCs in the inside air are enriched by means of active sampling on Tenax TA tubes, which are thermodesorbed. After internal standards have been added, analysis is carried out by capillary GC − MS. The hundred compounds on the chromatogram which have the most intense signals are identifi ed by retention index and mass spectrum (Wensing, Schulze and Salthammer, 2002 ), and are then semi - quantitatively evaluated with toluene as the reference substance. The toluene equivalents are summed, and this result serves as a semi - quantitative estimation of the total VOC concentration (TVOC).

7.3.3 Qualitative Determination of VOC s (Identifi cation)

Although samples are taken and prepared in the same way here as in the semi - quantitative analysis, in the evaluation of the complex spectra of the substances, the multi - dimensional GC technique (GCGC − MS) is employed. This consists of two series - connected GCs combined with an MS detector. As a rule, the process


is adjusted in such a way that polar compounds containing heteroatoms are extracted from the non - polar hydrocarbon matrix (which is of no interest) and identifi ed. The laborious work of identifi cation is facilitated by a special powerful database program that works on the principle of standardized retention time indices (Wensing et al ., 2002 ).

7.3.4 Identifi cation of SVOC s (Fogging Precipitate)

The samples are taken from the surface of the two cooled glass plates of a special apparatus (Figure 7.5 ) which is exposed to the air inside the car at the standardized interface. The glass plates (P) are mounted on both sides of the cooling body (K). The body is hollow inside with a coolant running through it. The temperature of the cooling body is registered by a rod - shaped measuring sensor which is intro-duced into the drilled hole (F) and reaches right into the center of the plate. While the samples are being taken, the vehicle is heated to increase emissions and thus to speed up the collection of substances. Conceptually, the procedure leans on the sample - taking technique of DIN 75201 (1992) . After the sample is taken, the fi lm which built up on the glass plates is washed off with organic solvent. The solution is then concentrated and analyzed by GC − MS.

7.3.5 Measurement of the Sum of Organic Substances ( Σ VOC )

For each experiment on a vehicle at the test stand, a FID is used for the sum determination of total organically fi xed carbon in the inside air. In this way it is possible to follow the relative dependence of the total concentration of VOCs on various infl uences, and at the same time the state of the inside air can be continu-ously documented.

Figure 7.5 Apparatus for the collection of SVOCs.


7.4 Quantitative and Qualitative Results from Brand New Cars

In an experiment lasting several years, six brand new cars were tested at the test stand for 40 days. For 8 hours each day the inside air was heated up to 65 ° C (arti-fi cial ageing). At the beginning and after 20 and 40 days the air was characterized by means of the standardized measurement procedures described above (T Ü V NORD, 1996 ).

As the sum of VOCs results (FID) in Figure 7.6 show, the measured values for concentrations and calculated ER decrease rapidly over time. The shape of the curve follows a simple exponential function. The calculated emission rate is directly linked to the source strength of the various interior components. The resulting concentration is also infl uenced by the air tightness of the car cabin. In accordance with the FID measurements, the decrease in the intensities of the individual signals and the clear change in their pattern can be seen in the VOC gas chromatogram (Figure 7.7 ). Calculation of the measured values of TVOC (toluene equivalents, TE) from the gas chromatogram of the six vehicles results in the following ranges:

Figure 7.6 Decrease in the sum of VOCs for six brand new cars during artifi cial ageing at 65 ° C.

7.4 Quantitative and Qualitative Results from Brand New Cars 157

Figure 7.7 Changes in the VOC gas chromatogram for brand new cars during artifi cial ageing.

• new condition: 7000 – 24 000 μ gTE/m 3 ; • after 20 days of ageing: 2500 – 10 000 μ gTE/m 3 ; • after 40 days of ageing: 1000 – 4500 μ gTE/m 3 .

Measured values of ER are given for a selection of VOCs in Table 7.1 . Generally, emission decreases from high measured values at the beginning to

lower ones subsequently. Because of these effects, the conclusion can be drawn that the processes are controlled by evaporation. Substances at the surface or in


Table 7.1 Emission rates of selected VOC s : arithmetical mean of six different new cars (T Ü V NORD, 1996 ) .

Substance New vehicle ER ( μ g/h)

After 20 days of ageing After 40 days of ageing

Benzene 24 22 27 Toluene 275 78 73 Ethylbenzene 434 38 18 m/p - Xylene 965 95 46 o - Xylene 272 37 20 Styrene 369 88 68 2 - Methoxyethanol 3.4 0.7 0.8 2 - Ethoxyethanol 2.6 0.7 0.7 2 - Butoxyethanol 68 12 5.6 2 - Ethoxyethyl acetate 4.4 1.4 1.3 2 - Butoxyethyl acetate 25 8.6 9.5 1 - Methoxypropyl acetate 104 24 6.8 Formaldehyde 40 52 43 Acetaldehyde 44 30 29 Propanal 15 9.5 13 n/iso - Butyraldehyde 19 7.2 7.2 Pentanal 29 14 16 Hexanal 42 25 13 Benzaldehyde 10 34 11 Acetone 361 143 116 Methylethyl ketone 85 13 7.5 Methylisobutyl ketone 755 78 47 Di - n - butyl phthalate 0.7 0.8 2.0 Dimethylamine 9.0 3.1 2.5 Diethylamine 9.6 8.1 5.7 Di - n - butylamine 54 14 59 N - nitrosodimethylamine 0.20 0.12 0.07

layers near the surface of the materials of the interior furnishings are desorbed rapidly. However, there are exceptions to these reactions, such as the aldehydes. Here, diffusion - controlled emissions can be assumed. In a slow and steady process, diffusions from the inside of the material take place which fade away only after a longer period of time. In both cases, textiles and textile (laminated) compound materials play an essential part as storage media and as a reversible interim store for chemical substances due to their strong sorption ability (Ehrler, Schreiber and Haller, 1994 ).

The apparatus described in Figure 7.5 was used for the accumulation of SVOCs. The accumulation capacity during the experiments amounted to 10 μ g of fi lm mass per glass plate per hour at a constant inside air temperature of 65 ° C. Examples of SVOCs found in fogging fi lms are listed below:

7.5 Emissions of Organophosphate Esters inside Automobiles 159

Hydrocarbons : alkanes, n - alkanes (CI3 - C32); branched alkanes Alcohols : 2 - ethylhexanol, octadecanol. Amines : 1,4 - diazabicyclo - 2,2,2 - octane; dicyclohexylamine, methyldicyclohexyl-

amine; dis - dimethylaminodiethyl ether; N,N - dimethylpentadecylamine. Amides : tridecylcycloacetamide; pentadecylcycloacetamide. Aromatic carbonic acid esters : dibutyl, dioctyl, di - (2 - ethylhexyl) phthalate; diocty-

ladipate, dioctylsebacate, trihexyltrimellitate. Fatty acids : lauric acid; stearic acid. Fatty acid esters : palmitic acid butyl ester; palmitic acid 2 - ethylhexyl ester; stearic

acid 2 - ethyhexyl ester; dibutyl adipate. Phenols : 2 - (1,1 - dimethylethyl) - phenol; 2,6 - di - tert - butyl - 4 - methylphenol ( BHT );

2,6 - di - tert - butyl - 4 - ethylphenol; 2,6 - di - tert - buyyl - 4 - methoxymethylphenol. Phosphates : tris - (2 - chloroethyl) phosphate; tris - (1 - chloro - iso - propyl) phosphate;

tris - (1,3 - dichloro - iso - propyl) phosphate. Other compounds : benzoic acid; 2 - ethylhexaneacid; erucic acid amide;

siloxanes.

7.5 Emissions of Organophosphate Esters inside Automobiles

In a special exposure study (Wensing, Pardemann and Schwampe, 2003 ) concern-ing SVOCs, the passenger compartments of eight new automobiles were tested on the automobile test stand for the emission of organophosphate esters both at room air temperature (RT; approx. 20 ° C) and in a heated - up state at 65 ° C. The collecting phase used for sampling the organophosphate esters was an XAD - 2 (500 mg) specially cleaned with upstream glass wool using various solvents (acetone, methanol, dichloromethane). A quantitative assessment (VDI 4301 - 5, 2009 ) of the organophosphates was carried out using solvent desorption (dichlo-romethane) and GC − MS evaluation (DB 1701, HP 5890/HP 5989A) using original standard solutions (Aldrich) and applying the internal standard method ( 13 C 12 - DDE) in SIM mode (VDI 4301 - 5, 2009 ). Table 7.2 shows the concentration ranges of the organophosphate esters that were measured under test stand conditions at RT and 65 ° C in the inside air of the eight different new vehicles from different manufacturers. The infl uence of the temperature on the release of the organo-phosphate esters used as fl ame retardants and/or plasticizers can be clearly seen here. Whereas at RT the individual compounds could either not be detected or were clearly below 1 μ g/m 3 , at 65 ° C comparably higher concentrations were measured.

For this reason the present study also carried out measurements with one car (vehicle V - X) under traveling conditions with regard to an exposure evaluation of organophosphate esters immediately subsequent to the test stand investigations. For results see Table 7.3 . In the test stand measurements at 65 ° C individual cases of concentration values were obtained. As expected, however, the present measure-ments under traveling conditions have clearly revealed that after a few minutes no


Table 7.2 Organophosphate esters, chemical names, synonyms and concentration ranges ( μ g/m 3 ) for RT and 65 ° C (Wensing, Pardemann and Schwampe, 2003 ) .

Chemical name Synonym Concentration, RT ( μ g/m 3 )

Concentration, 65 ° C ( μ g/m 3 )

Tributyl phosphate TBP < 0.01 – < 0.15 0.16 – 2.18 Tris(2 - butoxyethyl) phosphate TBEP < 0.01 – 0.05 < 0.05 – 0.69 Tris(2 - ethylhexyl) phosphate TEHP < 0.01 – 0.09 < 0.01 – 0.19 Triphenyl phosphate TPP < 0.01 – 0.09 < 0.01 – 2.23 Tricresyl phosphate TCP < 0.01 – < 0.05 < 0.01 – < 0.05 Tris(2 - chloroethyl) phosphate TCEP < 0.01 – < 0.05 < 0.01 – 0.10 Tris(chloropropyl) phosphate TCPP < 0.01 – 0.48 0.07 – 11.1 Tris(2,3 - dichloro - 1 - propyl)phosphate TDCPP < 0.01 – 0.20 0.07 – 8.64

Table 7.3 Organophosphate esters, synonyms and concentrations ( μ g/m 3 ) for vehicle V - X , test stand and traveling (Wensing, Pardemann and Schwampe, 2003 ) .

Synonym V - X V - X V - X V - X V - X #1 a #2 a #3 b , #4 c , #5 d , RT e 65 ° C e 65 ° C e 50 ° C e 40 ° C e ( μ g/m 3 ) ( μ g/m 3 ) ( μ g/m 3 ) ( μ g/m 3 ) ( μ g/m 3 )

TBP 0.02 1.83 0.49 < 0.53 < 0.34 TBEP < 0.01 0.26 < 0.39 < 0.53 < 0.34 TEHP 0.09 0.43 < 0.39 < 0.53 < 0.34 TPP < 0.01 0.22 < 0.39 < 0.53 < 0.34 TCP < 0.01 < 0.05 < 0.39 < 0.53 < 0.34 TCEP < 0.01 0.10 < 0.39 < 0.53 < 0.34 TCPP < 0.01 0.60 < 0.39 < 0.53 < 0.34 TDCPP 0.03 8.64 0.47 < 0.53 < 0.34

a Measurement on the test stand. b Measurement when traveling, immediately after the test stand measurement, duration

12 min. c Measurement when traveling, immediately following #3, duration 12 min. d Measurement when traveling, immediately following #4, duration 25 min.

relevant concentrations of organophosphate esters could be detected in the interior vehicle air due to the high air exchange rates which were present. As regards exposure the concentration values occurring during traveling can be evaluated on the basis of an assessment concept which was introduced some time ago: a German committee has developed a method of calculating inside air guideline values on the basis of toxicological data (Ad - hoc Working Group, 1996 ). Here Guideline II (RW II) defi nes a value which requires immediate action. Guideline I (RW I = RW II/10) is the concentration which is not to be exceeded for lifelong

References 161

exposure. For TCEP, Sagunski and Ro ß kamp (2002) have derived values of RW II = 50 μ g/m 3 and RW I = 5 μ g/m 3 . In view of the insuffi cient toxicological data available for other organophosphorus compounds the authors suggest RW I/II as sum values for TCEP, TCPP, TBP, TBEP TEHP and TPP concentrations.

Concentration values in the magnitude of RWII (50 μ g/m 3 ) were not achieved in any case in the present study even under the extreme climatic boundary condi-tions of the test stand investigation. The results for RT at the test stand (see Table 7.3 ) revealed only very low concentration values across the board, with individual maximum values in the magnitude of 1/10 RWI and are therefore regarded as irrelevant. In the test stand measurements at 65 ° C, individual cases of concentra-tion values in the magnitude of RWI (5 μ g/m 3 ) and above were obtained. As expected, however, the present measurements under traveling conditions (Table 7.3 ) clearly revealed that even in the ‘ hot ’ state in comparison with RWI no relevant concentrations of organophosphate esters could be detected in the interior vehicle air on account of the high air exchange rates present. The detection level lay in the magnitude of 1/10 RWI; even a life - long exposure to this value is not regarded as critical.

7.6 Conclusion

The existing test stand method and the standardized analytical process are power-ful routine methods for delivering a reliable characterization of emissions in automobile interiors. The effects on the IAQ of changes in the interior furnishing can be followed precisely. When combined with emission - chamber measurements of single parts of the interior, components can be constantly improved and the quality of the materials for interior furnishings secured (Bauhof et al ., 1996 ; VDA, 2005 ).

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Material and Indoor Odors and Odorants Florian Mayer , Klaus Breuer , and Klaus Sedlbauer

8

8.1 Introduction

Indoor odors and odorants are present in almost every indoor environment. If certain odor perceptions cannot be clearly attributed to common activities usually performed in (living) rooms, such as cleaning or preparing food, but are emanat-ing from various materials like building products, indoor materials or domestic appliances, odors often cause fears regarding health impairment due to emissions from materials, or they serve as an alleged reason for feelings of illness. Possible interrelations between odor perceptions, annoyance or nuisance caused by odors, individual associations and other phenomena such as the Sick Building Syndrome have been discussed for years (Fanger, 1987 ; Wargocki et al ., 1999 ; Wolkoff et al ., 2006 ). Other indoor environments are also attracting attention. The odor of a new car has an increasing impact on customer acceptance and therefore on brand and model selection. Car manufacturers have realized the sensibility of their customers regarding smell and the necessity of reducing and improving the odor of their products. The automotive industry has an increasing interest in investigating and optimizing the odor of the materials used in cars (Mayer and Breuer, 2004b ; Mair, Mayer and Breuer, 2005, 2006 ). Another important indoor environment is the cabin of an aircraft. Increasing numbers of passengers are traveling to more, and more remote, places and spend more time on board an aircraft. The quality of the cabin climate, especially the low humidity and also contamination with bioeffl u-ents, perfume, exhaust, toilet or galley odors, has become a topic of public interest (Lee et al ., 1999 ; During, 2005 ). This situation has urged researchers and R & D departments of industrial manufacturers to combine their efforts with the aim of investigating and evaluating materials regarding their odor properties (Breuer and Mayer, 1998 ). Two important objectives are given priority in this process, namely to develop reproducible methods to evaluate material odors, in order to be able to identify products that are ‘ peculiar ’ in terms of odor and, secondly, to improve the


165

166 8 Material and Indoor Odors and Odorants

quality of products by reducing their TVOC emissions (e.g., solvents) during their useful life.

But research over the past ten years reveals that in most cases there is no direct correlation between VOCs measured by routine emission tests and odor active compounds (Mayer and Breuer, 2000 ; Knudsen et al ., 1999 ; Salthammer et al ., 2004 ). Figure 8.1 shows an example of the time - dependent TVOC values of a fl oor-ing material in comparison to odor intensity scores (see later).

As a consequence, the reduction of the emissions of major components con-tributing to the TVOC value will not necessarily result in material odor reduction. Odors of building products, interior decoration materials or domestic appliances are more or less a random side effect of the manufacturing process, and the reasons for that presumably can be found in formulation ingredients or the processing conditions. Investigations on material - inherent causes regarding the chemical nature of odorants responsible for a specifi c material ’ s odor are very rare. However, the knowledge of these odor active substances, their chemical structure and properties are a prerequisite for conclusions on sources and favor-able processing conditions for odorant generation, and therewith for fi nding start-ing points for odor minimization, product improvement, consumer information and improved perceived indoor air quality. For example, knowing the chemical structure of material odorants might provide information about the origin of inevitable odorous emissions, which are not harmful to health but rather are trace amounts of odor active compounds that can be detected by the very sensi-tive human nose. Often, the detailed structure of such odorants is not known and is diffi cult to fi nd out, since routine methods for emission analyses like GC − MS are not usually selective or sensitive enough to detect many of the odorous compounds, which are quite often only contained in a material in very small concentrations in the ppb to ppt level range. Recently, increased attention

Figure 8.1 Non - correlation of TVOC emission and odor intensity of a fl ooring material.

8.2 Odor Evaluation 167

has been paid to secondary emissions from different sources, for instance, ozone reactions with indoor air contaminants and material emissions also called reactive chemistry, and blamed for bad indoor air quality (Knudsen et al ., 1999 ; Tamas et al ., 2006 ; Uhde and Salthammer, 2007 ; Weschler, 2004 ; Wolkoff, 1999 ; Wolkoff et al ., 2006 ) and material odor and the original, primary odor of a material has hardly been investigated. Still, it has been acknowledged that routine emission analysis is not suffi cient to address all the problems of sensory or other occupant health - related irritations because of the limitations of currently applied analytical routine methods.

This article will review the development and progress of analytical procedures of material odor evaluation, analysis and identifi cation during recent years and the knowledge gained about compounds responsible for odor in the area of material science.

8.2 Odor Evaluation

To localize and identify products and materials with an intense, unpleasant or unacceptable odor methods for a reliable, reproducible evaluation of odor is neces-sary, even mandatory.

8.2.1 Indoor Environments

Different methods are in use to evaluate air quality in rooms and buildings. One example is the collection of an air sample of the room to be evaluated in an odor-less container, for example a plastic bag, and subsequent evaluation by dynamic olfactometry in a laboratory according to DIN EN 13725 (DIN EN (2003) ). The sample air is continuously diluted in steps of a factor of two and evaluated by a trained panel of at least four, preferably more, subjects and the dilution at which 50% of the panel cannot smell anything results in an odor unit value per m 3 . Practical experience has revealed that this dilution method can only be applied to highly odorous samples. Small differences in odor intensity between two weakly odorous samples cannot be distinguished. Furthermore, sampling into a con-tainer, transportation of the sample from the sampling site to the evaluation labo-ratory might raise the question whether the air evaluated really represents the air sampled, considering possible contamination by the container, permeation of compounds through the container wall, and the time elapsed between sampling and evaluation, which can allow compound degradation. Therefore direct evalua-tion should be favored but might not always be practicable. On - site evaluation is performed either by a small trained panel or by a huge untrained panel. Different scales for this procedure are in use, usually intensity scales and acceptability scales (for untrained panels) or decipol scales for trained panels (Bluyssen et al ., 1989 ; Pejtersen et al ., 1990 ; Wargocki et al ., 1999 ).


8.2.2 Materials

In principle, the same methods can also be applied for product and material odor evaluation. For dynamic olfactometry the same restrictions are valid, this method is only useful for intensively odorous and distinctly differing samples; sampling of headspace air above a material into a container and transportation to the olfac-tometer unit for evaluation according to DIN EN 13725 (DIN EN (2003) ) might be considered questionable, so direct odor evaluation would be preferred (Saltham-mer et al ., 2004 ). Direct odor evaluation of materials by a sensory panel is per-formed by putting the sample into a chamber and by evaluating the air having been in contact with the material for a certain period of time. This can be done in a static experiment (no air exchange in the sample container enabling only one reliable evaluation) or in a dynamic experiment (with a continuous air fl ow through the container which enables frequent evaluations of the air fl owing through). The static approach, for example, is used by the German Automobile Association (VDA) for the odor evaluation of materials according to VDA Standard 270 (Verband der Automobilindustrie, 1992 ). The advantage of this method is the small effort regarding time, manpower and instrumentation and a high opera-tional capacity; the disadvantage is its limited reproducibility, reliability and con-sequently higher error rate, especially with fairly odorous materials around the given limit value between passing and rejected samples (Mair, Mayer and Breuer, 2005 ). The dynamic approach is widely used in Scandinavia, for example by the Danish Society of Indoor Climate for the Danish Indoor Climate Label. Here CLIMPAQ glass chambers (Gunnarsen, Nielsen and Wolkoff, 1994 ) with a defi ned airfl ow and a defi ned diffuser are used to evaluate the air fl owing through the chamber loaded with a certain building material (Danish Society of Indoor Climate, 2003 ; Knudsen, Valbj ø rn and Nielsen, 1998 ; Knudsen et al ., 1999 ). A similar method is the expansion of existing testing standards requiring emission test chambers such as DIN EN ISO 16000 - 9 (DIN EN ISO 2008 ) for the analyses of VOCs emitted by building products by an additional sensory evaluation using the dynamic air fl ow through the emission test chamber not only for chemical analy-ses sampling but also for odor evaluation of the chamber air (Breuer and Mayer, 1998 ). A specially designed diffuser (bottom diameter 24 mm, top diameter 80 mm, angle 8 ° , length 190 mm, Bluyssen (1990) is attached to the emission test chamber through which the air exiting the chamber is led at a fl ow rate of 3.2 m 3 /h. The quality of the chamber air is evaluated by the panel in an odorless atmosphere at defi ned temperature and humidity conditions (23 ° C, 50% r. h.). A scheme of the set - up for material odor evaluation is shown in Figure 8.2 .

8.2.3 Panels and Scales

The evaluation is performed either by a large untrained panel or by a smaller trained panel. As odor characteristics to be evaluated an intensity scale, a hedonic


Figure 8.3 Odor evaluation scales for (a) intensity (VDI 3882 - 1); (b) hedonic odor tone (VDI 3882 - 2); and (c) acceptance (Clausen et al ., 1996 ; Clausen, 2000 ) .

Figure 8.2 Experimental set - up for material odor evaluation using an emission test chamber.

scale and for an untrained, na ï ve panel an acceptability scale can be used. A ques-tionnaire design for this kind of evaluation is shown in Figure 8.3 . For scoring odor intensity the scale according to the German Guideline VDI 3882 Part 1 (VDI, 1992 , Figure 8.3 a) can be used. The Danish Society of Indoor Climate (2003) uses an intensity scale according to Yaglou, Riley and Coggins (1936) . A hedonic scale is suggested by the German Guideline VDI 3882, Part 2 (VDI, 1994 , Figure 8.3 b). The acceptability scale is based on Clausen et al . (1996) and Clausen (2000) , Figure 8.3 c; it is widely used and has proven suitability (Breuer and Mayer, 2003 ; Danish Society of Indoor Climate, 2003 ; Knudsen, Valbj ø rn and Nielsen, 1998 ; Knudsen et al ., 1999 ).

A trained panel instead evaluates perceived air quality by using the decipol scale according to Fanger (1988) . For this evaluation method, the panel members have been selected and trained according to a procedure described by Clausen et al .


(1996) . During evaluation four reference points are always offered representing defi ned decipol levels (2, 5, 10, 20 decipol), so called milestones. Each decipol level is created by a defi ned acetone concentration evaporated into air and provided to the subjects:

decipol ppm acetone−( )× =−0 84 0 22 1. . (Bluyssen, 1990 ).

One decipol is defi ned as the perceived air quality of a room at equilibrium state caused by the emissions (bio effl uents) of an average thermally balanced person at a ventilation rate of 10 l per second. The correlation between decipol levels and the percentage of dissatisfi ed has empirically been determined by Fanger (Figure 8.4 ). The percentage of dissatisfi ed can also be correlated to acceptability (Figure 8.5 ). Consequently acceptability scores (A) can directly be converted into decipol scores (C) by:

C

e

e

A

A= ⋅+

⋅⎛⎝⎜

⎞⎠⎟

−⎛⎝⎜

⎞⎠⎟

− − ⋅

− − ⋅

−

1121

100 5 980 18 5 28

0 18 5 28ln ., ,

, ,

44

The ideal number of untrained panelists recommended for such an investigation is between 20 and 40. Panelists can be selected and their suitability checked fol-lowing ISO 8586 - 1 (1993) . A simple test can be performed according to ISO 8587 (1988) before using them as subjects for odor evaluation as described by Clausen et al . (1996) and Wargocki et al . (2003) .

Figure 8.4 Correlation between perceived air quality C and percentage of dissatisfi ed PD.

C PD= ( ) −( )−112 5 98 4ln . Fanger (1988)


When using trained subjects the number can be reduced to between 4 and 12. An investigation on the possible errors when performing sensory evaluations by static or dynamic methods, by using trained or untrained subjects and the infl uence of different scales was performed by Mair et al . (2006) . Based on these and earlier results it is recommended that either 30 na ï ve subjects or 8 trained subjects should be used, (for intensity and decipol scale) and a dynamic set up to get standard deviations for the different evaluation scales and methods below 10%.

If the panelists have some experience in odor evaluation in addition they might be able to state the perceived odor quality (musty, smoky, fatty, fruity, rubber - like, etc.).

To evaluate building products Kasche et al . (2005a, 2005b) and M ü ller et al . (2005) have suggested a new scale, Perceived Intensity Π (PI) with the unit pi. 0 pi is equivalent to the odor threshold of a reference compound, for example, acetone at a concentration of 20 mg/m 3 air, at which 50% of a human panel can detect it. The scale is linear, n times the odor threshold is n − 1 pi. The evaluation of the PI of a material is done by a trained panel using a reference scale of for example, six different pi values represented by different reference compound concentrations in air. The trained panel compares the perceived intensity of the material with the intensity of different concentrations of the reference scale and gives a pi score. The need of a new scale and unit, its advantages over existing ones, its dissemina-tion and implementation in sensory evaluation procedures will be revealed in the future.

Figure 8.5 Correlation between acceptability A and percentage of dissatisfi ed PD.

PDe

e

A

A=

+⋅

− − ⋅

− − ⋅

0 18 5 28

0 18 5 281100

, ,

, , (Clausen et al ., 1996 )


8.3 Odor Analysis – Odorant Identifi cation

8.3.1 Methods

The method of choice for the analysis of odorants and subsequent identifi cation is gas chromatography - olfactometry ( GC − O ), a method which combines the sepa-ration capability of volatile compounds by GC with the selective and sensitive odor detector human nose (Fuller, Steltenkamp and Tisserand, 1964 ). A scheme of the set - up can be seen in Figure 8.6 .

This procedure allows the differentiation of odor active compounds from odor-less substances within a complex mixture of volatiles. For decades this procedure has been successfully applied for aroma analyses of foods (Grosch, 1993 ). The mixture of volatile compounds either collected in a purifi ed organic solvent extract or in a defi ned headspace volume is separated into its different components by means of GC and the effl uent gas fl ow at the end of the GC capillary column is split between a FID and an experienced test person ’ s nose. By sniffi ng the column effl uent, the human nose is able to perceive the odor active compounds contained in a complex mixture and the test person can mark the corresponding spot in the FID chromatogram recorded in parallel and attribute a certain odor quality. A sample GC − O chromatogram of a solvent extracted material is shown in Figure 8.7 .

Further, a stepwise dilution of the investigated mixture of volatile compounds permits identifi cation of the odorants that contribute most to the sample odor (Acree, Barnard and Cunningham, 1984 ; Holscher and Steinhart, 1992 ; Mayer and

Figure 8.6 Scheme of gas chromatography – olfactometry (GC - O).

8.3 Odor Analysis – Odorant Identifi cation 173

Breuer, 2000 ; Ullrich and Grosch, 1987 ), namely those odorants that are still per-ceptible at maximum dilution. This procedure should be performed on at least two GC columns of different polarity. The important odorants are subsequently identifi ed by performing large scale sample preparations applying solvent extrac-tion, odorant purifi cation by high vacuum distillation to separate volatile from nonvolatile compounds, column chromatography to separate compounds according to polarity and concentration steps like solvent evaporation using micro - distillation to obtain odorant samples for identifi cation experiments using multidimensional gas chromatography – mass spectrometry – olfactometry ( MDGC − MS − O , scheme see Figure 8.8 , Mayer and Breuer (2006) ).

MDGC − MS − O enables the transfer of part of the column effl uent of a gas chromatographically separated sample (column 1 in GC 1) via a cryotrap onto a second column of different polarity for further gas chromatographical separation (column 2 in GC − MS) to increase the chances of getting a clean mass spectrum of the target, often low concentrated, unknown compound without interference of other higher concentrated substances in the sample. The transfer from column 1 to column 2 is controlled by a specially designed valve switching system which either leads the column effl uent to the FID and sniffi ng port for olfactory detection of GC 1 or to a cryotrap cooled with liquid nitrogen at − 100 ° C. After collection of the target part of the effl uent the cryotrap is fl ash heated at 250 ° C and the sample fl ushed onto column 2 which at the end is connected to a mass selective detector and a sniffi ng port to monitor the isolation of an odorant (GC − MS).

Figure 8.7 Sample GC − O chromatogram of a solvent extracted material.


8.3.1.1 Sampling of Volatiles and Odorants from Indoor Environments For odor analysis of indoor air the most common method of sampling is trapping of sample air on suitable adsorbents like Tenax, Carbotrap, charcoal, etc. The samples can be thermally desorbed or eluted by a solvent. Thermal desorption allows only one - time analysis whereas elution of the adsorbents allows multiple analyses of the sample including dilution and identifi cation experiments. The limitation of air sampling on adsorbents is the small volume of air that can be trapped on commercially available traps containing only a few milligrams of adsorptive material and allowing trapping of only 2 – 3 l of air at most without risking breakthrough of compounds. It is certainly possible to custom - make adsorbent traps using several grams of adsorptive material, but this is cost - intensive and huge traps are not very easy to handle. Volatiles can also be trapped by sucking air through one, preferably more, cold traps fi lled with a low boiling solvent (e.g., diethyl ether cooled in a dry ice/acetone bath at ca. − 80 ° C). That way several 1000 l of air can be sampled. The solvent in the traps is pooled, separated from water (frozen air humidity) in a separation funnel, dried by adding a desiccant (e.g., sodium sulfate), fi ltered and concentrated to a very small volume (0.1 – 0.5 ml) which

Figure 8.8 Scheme of multidimensional gas chromatography – mass spectrometry – olfactometry (MDGC − MS − O) for the identifi cation of trace concentrations of odorants.


can either be analyzed directly or further be purifi ed and separated into differently polar sub - fraction by column chromatography on silica gel (Mayer and Breuer, 2000, 2006 ). The big disadvantage of this method is that volatile compounds with a boiling point below or around that of the solvent used are lost or discriminated. For those compounds adsorbent trapping can be used complementarily.

8.3.1.2 Sampling of Volatiles and Isolation of Odorants from Materials For odorant analysis of materials a fi rst approach could be the analysis of the vola-tiles emitted into air by headspace sampling. A sample of the material is placed into a suitable glass container, equilibrated at room or elevated temperature, 80 ° C for instance, and then a sample is drawn by collecting air on a suitable adsorbent trap. The presence of many highly concentrated non - odorous compounds and the usual small concentration of odorants in a headspace sample makes odorant identifi cation very hard. Therefore the extraction of odorous material samples by a suitable solvent like diethyl ether, pentane or dichloromethane is often necessary to gain high enough odorant concentrations for identifi cation. The most suitable solvent for extraction of the odorants must be found in a small - scale pre - test. Depending on the odor intensity of the product, extraction is performed with up to several hundred grams of the material and up to 1000 ml of solvent (Mayer and Breuer, 2006 ). Small pieces or ground powder of the material are shaken or stirred with the solvent for some hours. The solvent is fi ltered and the extraction is repeated twice with fresh solvent. The solvent extracts are pooled. Since by solvent extraction many nonvolatile compounds are extracted that disturb GC analysis, the volatiles have to be isolated. This is performed by high vacuum distillation or by applying solvent - assisted fl avor evaporation ( SAFE ) using an apparatus designed by Engel, Bahr and Schieberle (1999) . The concentrated distillate is used for analy-sis or further purifi cation.

8.3.1.3 Identifi cation The important odorants which have been localized by GC − O and dilution analyses are identifi ed by GC − MS analyses of headspace samples or, for very low concen-trated odorants necessary, after solvent extraction, vacuum distillation/ SAFE and column chromatography by MDGC - MS.

Criteria for an unambiguous identifi cation of an odorant are:

• its odor quality

• its retention index (relative retention time compared to n - alkanes according to van den Dool and Kratz (1963) on at least two GC columns of different polarity;

• its mass spectrum;

• in accordance with the authentic reference compound.

Some reference odorants are commercially available; some have to be synth-esized in the laboratory. Sometimes or often, the concentration of an important odorant in a sample might be so low, that even after enrichment and purifi cation,


the amount is too small to obtain a meaningful MS, then the remaining criteria have to be in accordance with the reference.

Because of this complicated and time - consuming procedure attempts have been made to combine GC − O directly with GC − MS, even then it was diffi cult to iden-tify odorants directly, because the concentrations are usually too small and odorant peaks are hidden underneath highly concentrated non - odorous compound peaks. Clausen et al . (2005) investigated linseed oil and linseed containing fl oor oils and listed 139 odorants, but were only able to identify 45 odorants by GC − MS library search, leaving 76 unknown odorants and (at least) 18 questionable identifi cations, a big problem being co - elution of odorants with highly concentrated non - odorous compounds overlapping and dominating any MS obtained. Without huge sample preparations, enrichment and purifi cation steps, dilution experiments, analysis on at least two different GC columns of different polarity, knowledge on hundreds of known odorants and authentic reference compounds, the chances of unambigu-ous odorant identifi cation are quite small.

Still and so far, the analysis of the most important odorous substances in materi-als (unfortunately not always) by means of GC − O has revealed much useful knowledge on material odorants, possible sources and starting points for odor minimization and product improvement (Mayer and Breuer, 2004a, 2004b, 2005, 2006 ).

8.3.2 Examples

Some examples of odorants described in the literature which can be detected indoors or might emit from different products are given below. Odorants emitted by micro - organisms, which under certain favorable conditions can also occur indoors, are not considered, but might also contribute to indoor odor annoyance.

8.3.2.1 Cleaning Products, Detergents, Air Fresheners Many products contain perfumes and odorants added to cover up bad odors in homes and to suggest cleanness and freshness. Common examples of odor active compounds added are citronellol, geraniol, linalool, α - terpineol, citronellal, geranial, neral, linalyl acetate, camphene, limonene, β - myrcene, α - phellandrene, α - pinene, β - pinene, α - terpinolene, α - terpinene, octanal, nonanal, decanal (Salt-hammer, 1999b ) (see Chapter 15 ). Terpenes might be responsible for adverse health effects and might cause allergic reactions (Jensen et al ., 2001 ; Nazaroff and Weschler, 2004 ). It has been discussed that terpenes containing double bonds easily react with ozone, hydroxyl and nitrate radicals to form plenty of reaction products, so called secondary emissions, which can be odor active as well (Nazaroff and Weschler, 2004 ).

8.3.2.2 Carpets Important emissions from new carpets with a styrene - butadiene rubber latex backing adhesive are styrene, 4 - phenyl cyclohexene and 4 - vinyl cyclohexene. These


compounds, especially 4 - phenylcyclohexene have been held responsible for the typical ‘ new carpet ’ odor (Weschler, 1992 ), but this is not necessarily the case. Carpets also emit small amounts of some saturated C 6 – C 12 aldehydes. Their amount increases when the carpet is exposed to small amounts of ozone (Weschler, 1992 ), but also the amount of unsaturated aldehydes like 2 - nonenal grows (Morrison and Nazaroff, 2002 ). Saturated as well as unsaturated aldehydes are oxidation products of unsaturated fatty acids like oleic and linoleic acid. Oxidation agents can be oxygen (together with light) (Salthammer, Schwarz and Fuhrmann, 1999 ) as well as ozone (Thornberry and Abbatt, 2004 ).

8.3.2.3 Adhesives Oil - containing adhesives might emit fatty acid oxidation products like saturated and unsaturated aldehydes which can contribute to odor (Wilke, Jann and Br ö dner, 2004 ). Adhesives on a phenol resin base have been found responsible for odor annoyance in several offi ce buildings in former East Berlin. Alkyl - substituted phenols, methyl, dimethyl and ethyl phenols, some of which have very low odor thresholds, have been detected in indoor air as well as in different fl oor samples and were most likely responsible for the off - odor (Kirchner and Pernak, 2004 ).

8.3.2.4 Rubber Materials Used for Sealings, Floorings, Insulations Many of the odorous emissions from rubber and thermoplastic vulcanizate s ( TPV s) are primary and secondary products of the vulcanization reaction during which crosslinking of elastomers is achieved by addition of peroxides or, more frequently, sulfur together with nitrogen and/or sulfur - containing accelerators like benzothiazol derivatives, and heating up to 200 ° C for a few minutes. Much research on rubber emissions has been performed by the DIK (German Institute for Rubber Technology); examples are given by Giese (2005) . Compounds emitted by rubber are solvents and hydrocarbons, aldehydes and ketones, but also esters and phenolic compounds, hydrogen sulfi de, carbon disulfi de, and especially amines and benzothiazol derivatives which are mostly responsible for the typical odor of rubber. A rubber fl ooring investigated by Wilke, Jann and Br ö dner (2004) during a period of 28 days emitted at the end of this period mostly benzothiazol and some traces of naphthalene, styrene and benzaldehyde. The odor of a butadi-ene - styrene rubber used as insulation material was predominated by dimethyl disulfi de, ethyl methylpropanoate, ethyl 2 - methylbutanoate, ethyl pentanoate, 4 - vinyl cyclohexene and p - and m - cresol (Mayer, Breuer and Mayer, 2000 ).

8.3.2.5 Wood The odor of wood is mostly caused by the terpenes contained in the resin. Differ-ences in the composition of several kinds of resin have been investigated (Moyler and Clery, 1997 ). There is also some data available on the emissions of treated wood, wood products and furniture (Baumann, Batterman and Zhang, 1999 ; Risholm - Sundmann et al ., 1998 ; Salthammer, 1999a ; Salthammer, Schwarz and Fuhrmann, 1999 ). Besides terpenes like α - pinene, β - pinene, Δ 3 - carene, longi-folene, β - phellandrene, camphene, myrcene, carvone, limonene, caryophyllene,


borneol, camphor, fenchol, fenchone, p - cymene, compounds frequently found are aldehydes (C1 – C11 alkanals, C5 – C11 2 - alkenals and C7, C9 and C10 2,4 - alkadien-als) and solvents (e.g., toluene, xylenes, acetone, acetic acid, C3 – C5 alcohols, butyl acetate, 2 - butanone). Some photo - initiator fragments like benzaldehyde and ace-tophenone can also contribute to the odor of coated wood. All these fi ndings are based on emission analysis. Wood has hardly been analyzed by GC − O to look for low concentrated odorants and their contribution to the overall odor of different kinds of wood. One exception is the odor of oak which has been analyzed very thoroughly, since oak is often used as storage container for alcoholic beverages like wine, cognac, brandy, whiskey and odor and taste active compounds migrate from oak into the beverages and contribute to their fl avor (Flak et al ., 2004 ). Data on other kinds of wood are rare.

Gunschera et al . (2004) identifi ed odor active chloroanisols as degradation prod-ucts of pentachlorophenol used in the 1960s and 1970s in Germany for wood protection from fungi. Especially 2,3,4,6 - tetrachloroanisol is thought to be responsible for a musty odor in several different frame - houses. Micro - organisms presumably metabolize pentachloro - phenol via 2,3,4,6 - tetrachlorophenol to 2,3,4,6 - tetrachloroanisol. The odor thresholds of chloroanisols are in the ppt range.

8.3.2.6 Wood - Based Flooring Materials GC − O experiments of a waxed and oiled parquet fl oor revealed acetaldehyde, methyl propanal, 3 - methyl butanal, pentanal, hexanal, heptanal, α - pinene, 1 - octen - 3 - one, β - pinene, octanal, 1 - nonen - 3 - one, nonanal, Z - 2 - nonenal and E - 2 nonenal as the most important odorants (Mayer, Breuer and Mayer, 2000 ). In contrast the less intense odor of varnished parquet fl oor was mostly caused by acetaldehyde, butanal, 3 - methyl butanal, pentanal, hexanal, α - pinene, benzaldehyde, 1 - octen - 3 - one, β - pinene and acetophenone.

Clausen et al . (2005) found many similarities between odorants emitted from linseed oil as well as from a fl oor oil made of this linseed oil, concluding that the odorants of the linseed oil are also responsible for the odor of the fl oor oil. Of the 139 listed perceived odorants only 45 were identifi ed by GC − MS library search and retention characteristics. Important odorants with a high detection frequency were acetaldehyde, propanal, butanal, pentanal, 2 - pentenal, hexanal, 2 - hexenal, heptanal, 2 - heptenal, 2,4 - heptadienal, octanal, 2 - octenal, nonanal, 2 - nonenal, 2 - decenal, benzaldehyde, 1 - penten - 3 - one, 1 - penten - 3 - ol, pentyl oxiran, acetic acid, propionic acid, butanoic acid, pentanoic acid, hexanoic acid, octanoic acid.

8.3.2.7 Linoleum Linoleum was analyzed by Jensen, Wolkoff and Wilkins (1995, 1996) using GC − O and GC − MS. The most important odorants identifi ed were: butanal, hexanal, heptanal, octanal, nonanal, decanal, acetic acid, propionic acid, butyric acid, hexa-noic acid, toluene, 2 - pentylfuran. The authors state that they might have over-looked some unsaturated aldehydes with very low odor thresholds. They were able to closely simulate the odor of linoleum by mixing the compounds according to


the headspace data obtained by emission measurements. Our own GC − O experi-ments of a linoleum fl ooring revealed in addition acetaldehyde, propanal, 2,3 - butandione, 3 - methyl butanal, 1 - penten - 3 - one, pentanal, 1 - octen - 3 - one, E - 2 - octenal, Z - 2 - nonenal, E - 2 - nonenal, E,E - 2,4 - nonadienal, E - 2 - decenal and E,E - 2,4 - decadienal as important contributors to linoleum odor. 2 - Pentylfuran was not perceived as an odorant (unpublished results).

8.3.2.8 Gypsum - Based Products Gypsum usually has an unobtrusive rather pleasant smell. Because of its salt structure (CaSO 4 ) it would be expected to be odorless, but empirically it is well - known that products with a gypsum basis develop a typical odor described as gypsum - like, slightly milky and chalk - like.

Gypsum is for example, used as drywall, stucco or constituent in mortar. Burdack - Freitag, Mayer and Breuer (2008) found as odorants aldehydes (3 - methylbutanal, hexanal, octanal, nonanal, (Z) - 2 - nonenal), ketones (2,3 - butandi-one, 1 - hexen - 3 - one, 2 - octanone, 1 - octen - 3 - one), acetic acid, 2 - acetyl - 1 - pyrroline and traces of some sulfur containing odorants, sulfi des (diethyl sulfi de, 2 - (ethylthio) - propane, butyl ethyl sulfi de), disulfi des (methyl ethyl disulfi de, ethyl isopropyl disulfi de, diisopropyl disulfi de, isobutyl isopropyl disulfi de), thiols (methanethiol, propanethiol, 3 - methyl - 2 - butene - 1 - thiol) and aromatic sulfur compounds (2 - methyl - 3 - furanthiol, diethylthiophene) at very small concentrations ( < 1 μ g/m 3 ). Gypsum deposits naturally contain small amounts of sulfur which during produc-tion (grinding, burning) can react with organic compounds, for example, from beater additives, to form organic sulfur compounds. Depending on the raw mate-rial or processing conditions sulfur compounds emitted at small concentrations give gypsum products a pleasant, milky, gypsum - like odor; at too high concentra-tions they cause unpleasant sulfurous off - odors.

8.3.2.9 Plastics The odor of plastics has hardly been investigated. The focus of most investigations is emission measurement of VOCs detectable by routine GC − MS analysis because emission limits have to be met by manufacturers to get approval for use of their material by authorities or by other companies which use the initial material for further processing. Examples of plastic off - odor analysis have been reported on certain occasions especially for food packaging materials regarding migration from the packaging into the food (Skjevrak et al ., 2003 ). Other published data on the odor of plastic materials are rare for confi dentiality reasons. A few selected examples of odorants found in certain plastic materials are mentioned below:

Thermoplastic Polyolefi n (TPO) One batch of a thermoplastic polyolefi n had a roasty off - odor. The important odorants 2,3 - butandione, 1 - hexen - 3 - one, methional (3 - methylthiopropanal), Z - 2 - nonenal, E - 2 - nonenal, 1 - octen - 3 - one, octanal, E,E - 2,4 - nonadienal, E,E - 2,4 - decadienal, and as the most important off - odorant 2 - acetyl - 1 - pyrroline, could be identifi ed (Mayer and Breuer, 2004b, 2006 ).


Polyphenylenoxide (PPO) Substituted phenols are used as monomers for the production of polyphenylenoxide s, ( PPO s) so they as well as phenolic degradation products can be found as emitted odor active compounds. In one case the odor of a PPO was predominantly caused by 2,6 - dimethylphenol and trimethylanisol as well as by a tentatively identifi ed substituted methoxypyrazine (Mayer and Breuer, 2004a ). Another potent odorant derived from higher molecular phenolic com-pounds, antioxidants for example, by the infl uence of heat ( > 200 ° C) and pressure is guaiacol (2 - methoxyphenol) (Mayer and Breuer, 2006 ).

Polyurethanes (PU) The odor of many polyurethanes is caused by fi shy smelling low molecular weight amines, mostly methylamines like trimethylamine, degrada-tion products of the amine catalysts used during production. In some cases after evaporation of the fi shy smelling amines, other odor qualities remain. One notice-able odor of PU samples was earthy, nutty, which can often be attributed to sub-stituted pyrazines, formed by condensation of two α - aminoketones, subsequent oxidation or alkylation by aldehydes. One example is 2 - ethyl - 3,5 - dimethylpyrazine (Mayer and Breuer, 2006 ).

8.3.2.10 Electronic Devices Emissions from electronic devices have been investigated by M ö ller et al . (2003) ; Wensing (2004) and their infl uence on the sensory quality of offi ce air by Wargocki et al . (2003) ; Bak ó - Bir ó et al . (2004) . Important compounds emitted from electronic offi ce equipment like television sets, video recorders and computer screens which might contribute to odor are phenols and cresols emitted from circuit boards. New TV sets and cathode - ray - tube computer monitors have a negative infl uence on perceived air quality in offi ces. Aldehydes have been a focus of some investiga-tions, but the concentrations of the aldehydes found were below their reported odor thresholds. The fi nding that measured concentrations of emitted compounds cannot explain the negative effect on perceived air quality and the assumption that compounds not detected by routine emission analysis are responsible for these effects, once again reveals the inadequacy of currently applied routine emission and indoor air quality analysis and confi rms that the human sense often is more sensitive than chemical analysis. The term ‘ stealth chemicals ’ attributed by Bak ó - Bir ó et al . (2004) to these non - detected compounds is misleading because applica-tion of suitable methods like GC − O with a human sensor and large scale sampling permit detection and identifi cation of these compounds which are, granted, far from routine because of the time and effort needed for this kind of analysis.

More information about emission testing of electronic devices can be found in Chapter 17 .

8.3.3 Odorants and Odor Thresholds

The question whether a compound is odor active depends on its concentration and its odor threshold. Many compounds are odorous but some of them only if


Table 8.1 Examples of odor thresholds of some important material odorants mentioned.

Odorant Odor quality Odor threshold (ng/l air) a

Toluene Solvent - like 5900 b Limonene Citrus - like 135 α - pinene Terpeny, woody 18 Octanal Soapy, citrus - like 6 E - 2 - nonenal Fatty 0.1 E,E - 2,4 - decadienal Fatty 0.04 1 - octen - 3 - one Mushroom like 0.03 Ethyl 2 - methyl butanoate Fruity 0.06 Phenol Phenolic 430 b 2 - methylphenol (o - cresol) Phenolic 0.3 2,6 - dimethylphenol Phenolic, bitter 4 b 2 - methoxyphenol (guaiacol) Smoky 0.1 Methanethiol Rotten 2 b 2 - methyl - 3 - furanthiol Meaty 0.001 Trimethyl amine Fishy 2.5 2 - ethyl - 3,5 - dimethyl - pyrazine Earthy 0.007 2 - acetyl - 1 - pyrroline Roasty 0.02

a Rychlik, Schieberle and Grosch (1998) . b Devos et al . (1990) .

the concentration is very high, because their odor thresholds are very high as well, toluene for example. But there are also compounds that have very low odor thresh-olds, in the ppb and ppt level range. Consequently, very small amounts of these compounds can contribute to odor. Examples of odor thresholds of some material odorants mentioned above are given in Table 8.1 .

If compounds with very low odor thresholds and very small concentrations contribute to a material ’ s odor their detection can be very challenging, especially when only applying routine emission measurements like GC − MS. Such com-pounds will easily be overlooked, for their detection GC − O can often be the only choice, but so far this method is seldom used in material analysis. Instead con-centrations determined by emission measurements are compared with published odor thresholds to decide whether a compound might contribute to the odor or not. One problem is that published odor thresholds can differ quite a lot, even by several orders of magnitude (van Gemert, 2003 ). The value depends on the method and the panel but also on the purity of the compound used for threshold determi-nation (if small impurities of a substance with a low odor threshold were present in a sample the odor threshold determined would have been too low!). Many factors infl uence odor threshold determination, therefore many published values are questionable and they are hard to rely on. Some authors (Knudsen et al ., 1999 ; Wolkoff, 1999 ; Wolkoff et al ., 2006 ) assume that many of the odor thresholds reported in the literature are actually much lower, because if they compare con-centrations of compounds emitted and measured with odor thresholds published,


most materials should be odorless after a while. Often materials still have a distinct odor despite measured concentrations being below reported odor thresholds. The answer might be that many compounds which are responsible for odor have never been measured because their concentrations in a material are so low, below the detection limit of routine emission measurements, and their odor thresholds are lower than many of well known compounds emitted and measured for decades. This fact is not a matter of some odor thresholds previously published being too high but rather based on the lack of proper analysis methods used, which are not necessarily suitable for odorant analysis. GC − O of an odorous material sample will give reliable information on the odor potency of single odorants if the person evaluating the column effl uent has an average sense of smell (no anosmia) and some experience. Compounds with very small concentrations and very low odor thresholds will be detected by this method. Still, unambiguous identifi cation of such compounds is the next challenge.

8.3.4 Application of the Combination of Odor Evaluation and Odor Analysis for Product Optimization

The information gained by application of odor evaluation and odor analysis of indoor environment and material samples not only serves the purpose of knowl-edge accumulation, but can also be used as a tool for air quality and product improvement. The combination of the two methods has the potential of a system-atic approach to developing odor optimized technical materials. One example is given here.

A PPO that emitted odorants, substituted phenols like those mentioned above, was evaluated by a panel of 30 na ï ve subjects using the intensity and hedonic scale according to VDI 3882 Part 1 and 2 (1992, 1994) and the acceptance scale accord-ing to Clausen et al . (1996) . The result is shown in Table 8.2 (acceptance was converted into decipol, Figures 8.4 and 8.5 ). After identifi cation of the odorants and, based on identifi cation of source and generation mechanisms, modifi cation of formulation and processing conditions an odor optimized PPO could be pro-duced. GC − O of this PPO revealed that the compounds responsible for the intense odor of the previous product has either been removed or signifi cantly reduced. Odor evaluation of the new product by the panel using the same scales resulted in much better scores as shown in Table 8.2 .

Table 8.2 Result of the odor evaluation of an odorous and odor improved PPO .

Odor intensity Hedonic odor tone Perceived air quality (decipol)

Odorous PPO 3.8 − 2.1 22.6 Odor optimized PPO 2.6 − 0.6 12.9

8.4 Conclusion and Outlook 183

Lower values for all sensory characteristics evaluated prove the successful opti-mization of the material and the product subsequently made from it.

8.4 Conclusion and Outlook

Because of the frequent non - correlation of VOC emissions and the odor of a tech-nical product, quality evaluation of a product for indoor application should be based not only on chemical measurements but also on sensory experiments with material odor evaluation (Knudsen et al ., 1999 ; Salthammer et al ., 2004 ) and, if necessary, detailed odor analysis. For odorant detection the human sense of smell is much more sensitive than currently available analytical detectors. In indoor environments odor evaluation and odor analysis is a useful tool for the search of responsible sources of odor. For materials, the localization of responsible odorants and identifi cation of their chemical structure allow conclusions to be drawn on the sources and conditions for their formation, whether that be a certain technical ingredient or a particular processing step. The combination of odor evaluation and odor analysis is a very useful, even necessary, tool to monitor and improve material and product quality regarding odor. If a certain material spreads an intense, unusual, or unpleasant odor, odor evaluation and odor analysis can provide valu-able information in addition to any routine emission measurement data. After identifi cation such odorants could be evaluated with regard to health aspects like sensory irritations and allergies. Since material odors are probably inevitable in some cases, this method might also open up new ways of informing consumers. Some intensively odor active compounds detectable by the sensitive human nose might only be emitted at very small concentrations and are probably well - known odorants from familiar non - material sources, food products, for instance, which will be very useful to inform and calm consumers about the odorous side effect of material production.

The information on odorous substances gained can also be used in investiga-tions on the effi ciency of odor minimizing products like fi lters, adsorbents, cata-lytic active materials which are expected to retain, bind or convert odorants from air to improve air quality. Odor evaluation and odor analysis by means of GC − O can help in testing the odor removal capacity of such products, for instance by addition of very small amounts of very potent odorants to see how much of the compound is held back and whether traces still go through and still are perceptible downstream of the device. When looking for examples at (photo - ) catalytic active products applied to purify air by conversion of organic contaminants to carbon dioxide and water, eventually, GC − O can help to reveal whether unobtrusive odor-less organic compounds are converted into much more disturbing compounds with low odor threshold by partial, incomplete oxidation. These methods are used to test the effi ciency of air fi lter and purifi cation systems for automotive and air-craft industry (Nurcombe, 2005 ).


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Mair , S. , Witek , W. , Dahn , U. , Mayer , F. and Breuer , K. ( 2006 ) Zur Pr ä zision humanolfaktometrischer Messungen . Gefahrstoffe – Reinhaltung der Luft , 66/3 , 94 – 8 .

Mayer , F. and Breuer , K. ( 2000 ) Geruchsstoffe von Bauprodukten in Innenr ä umen – Gaschromatographisch - olfaktometrische Untersuchung des Materialgeruchs eines Parkettbodens . Bauphysik , 22 ( 2 ), 96 – 100 .

Mayer , F. and Breuer , K. ( 2004a ) Kontrollierte Humanolfaktometrie in Kombination mit Geruchsanalytik als Entwicklungsinstrumentarium im Automobilbau . Proceedings of the 6th Workshop ‘ Odour and Emissions of Plastic Materials ’ , Kassel, Germany , pp. 15 - 1 – 15 - 12 .

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Mayer , F. , Breuer , K. and Mayer , E. ( 2000 ) Determination of Odoractive Volatiles Emitted by Building Materials by a New Method using Gas Chromatography - Olfactometry . Proceedings of Healthy Buildings 2000 , Vol. 4 ‘ Materials, Design and Construction ’ , Helsinki, Finland , pp. 119 – 24 .

Morrison , G.C. and Nazaroff , W.W. ( 2002 ) Ozone interactions with carpet: secondary emissions of aldehydes . Environmental Science and Technology , 36/10 , 2185 – 92 .

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Evaluation of Indoor Air Contamination by Means of Reference and Guide Values: The G erman Approach Birger Heinzow , and Helmut Sagunski

9

9.1 Introduction

Health and well - being can be negatively affected by pollutants in the indoor air. In response to reported feelings and symptoms of ill - health and/or the perception of odors in indoor areas, indoor air measurements are frequently ordered in both the private and public domains. Alongside offi cial bodies a number of private appraisers, institutes and laboratories are active in this fi eld. In practice it appears that comparable procedures and standards are not always applied when conduct-ing and evaluating measurements of this kind. In order to harmonize procedures in Germany, recommendations on the evaluation of indoor air contamination by means of reference and guide values have been prepared by the Ad - hoc Working Group of the Indoor Air Hygiene Commission of the German Federal Environ-ment Agency and of the Supreme State Health Authorities (IRK/AOLG Ad - hoc working group). According to a resolution of the Conference of Health Ministers, the responsibility of this committee is to lay down toxicologically based guide values in Germany (Ad - hoc AG, 1996 ). The aim is to minimize possible uncertain-ties and divergences in evaluation in the measurement and evaluation of indoor air quality and thus prevent the ensuing possible annoyance of affected parties and consequent disputes. These recommendations focus on the evaluation of VOCs but also apply to other indoor substances of interest for which guide or reference values are available. The evaluation of measurement results for indoor air is in principle based on an evaluation hierarchy:

i) as a health evaluation makes use of toxicologically derived guide values for individual substances or groups of substances;

ii) as a comparative evaluation is guided by statistical values, for example, the reference values of individual substances and the total VOC s value ( TVOC value).


189

190 9 Evaluation of Indoor Air Contamination by Means of Reference and Guide Values

These recommendations provide a practice - related procedure which refl ects the current regulatory state of discussion in Germany and which should be used as a mandatory and differentiated evaluation code for the public sector and also as a recommendation for the interior of private buildings. They are primarily directed at persons employed in public bodies (e.g., public health and environmental authorities), at appraisers and at measurement institutes concerned with health questions relating to indoor air quality.

9.2 Defi nition of Terms

9.2.1 Indoor Environment

On the basis of the defi nition of the Sachverst ä ndigenrat f ü r Umweltfragen (Council of Environmental Advisors) (SRU, 1987 ), of guideline VDI 4300 Part 1 (VDI 4300, 1995 ) and of ISO 16000 - 1 (DIN EN ISO 16000 - 1, 2006 ) the following are defi ned as indoor environments:

• private living and recreation spaces such as living rooms, bedrooms, bathrooms, kitchens, hobby rooms, sports rooms and basements;

• spaces in public buildings (e.g., schools, kindergartens, youth centers, hospitals, sports halls, libraries, restaurants and other public event venues);

• work areas and workplaces in buildings which, as regards air pollutants, are not subject to the provisions of legislation on hazardous substances (particularly regarding workplace limit values);

• passenger compartments in motor vehicles and public transportation.

For work areas (rooms inside buildings in which workplaces have been perma-nently installed) the requirements relating to workplaces as specifi ed in the Work-places Ordinance must be observed. Basically rooms in workplaces (such as offi ce rooms, for example) are deemed to be indoor environments in the aforementioned sense when the air pollutants found there are not used as working materials or when a working material from an area subject to hazardous substances legislation crosses into these rooms (Welzbacher, 1999 ).

9.2.2 Utilization Cycle

The term utilization cycle is defi ned as the period of time between two instances of ventilation, for example, lesson period(s) or the time of occupancy between two instances of ventilation at lesson breaks or recess.

9.3 Values for Evaluating the Indoor Air Quality 191

9.2.3 Volatile Organic Compounds ( VOC s )

In accordance with international recommendations the term volatile organic com-pounds (VOCs) refers to organochemical compounds in the boiling point range of approx. 50 to 260 ° C (WHO, 1989 ); this range is nearly identical to the defi ni-tions of ECA (1997) and AgBB (2008) . According to ECA and AgBB, VOCs are organic compounds which in a nonpolar column can be detected analytically in the elution range between n - hexane and n - hexadecane (AgBB, 2008 ; Seifert, 1999 ). Within the context of the TVOC concept, VOCs can be individual identifi ed and unidentifi ed substances. TVOC designates the totality of volatile organic com-pounds as the sum of instrumental responses for individual compounds between n - hexane and n - hexadecane (DIN ISO 16000 - 6, 2004 ). The identifi ed substances have to be quantifi ed substance - specifi cally with the aid of individual standards, the unidentifi ed ( ‘ unknown ’ ) substances in each case being quantifi ed as a toluene equivalent. Since various substances elute in the region between n - hexane and n - hexadecane depending on gas - chromatographic conditions (temperature program, column, and so on), in TVOC calculation it is recommended, as regards demarcation from very volatile organic compound s ( VVOC s), that the AgBB ’ s selection of substances for VOCs be followed. Semi - volatile organic compound s ( SVOC s) are organic compounds which are found in the retention region above n - hexadecane up to C 22 , while VVOCs are those occurring below n - hexane.

9.3 Values for Evaluating the Indoor Air Quality

Neither in Germany nor in Europe are there any regulations with broad legally binding force for quality requirements relating to indoor air. On the other hand a number of evaluation values do exist which have different names depending on the particular author (e.g., ‘ guide values ’ , ‘ orientation values ’ , ‘ target values ’ , ‘ precautionary values ’ , ‘ attention values ’ ) and which vary considerably in their technical derivation and legal signifi cance (Fromme, 2003 ). Toxicologically based values should always be distinguished from statistically defi ned reference or background values.

9.3.1 Toxicologically Based Values

Guide values are characterized by being based on relevant fi ndings regarding toxic effects and the dose – effect relationships of the substance in question; to protect sensitive groups of the population they often include safety margins. Precaution-ary values are as a rule set at a specifi c level below toxicologically based values and are intended to keep exposures and risks low and secure health protection in the long term.


9.3.2 Statistically Defi ned Values

Reference values refl ect the general exposure to a substance ( ‘ background expo-sure ’ ) and provide no information about health risk. According to publications of the international societies IFCC and IUPAC (Solberg, 1987 ; Poulsen, Holst and Christensen, 1997 ) a reference value for a chemical substance in an environmental medium is a value which has been derived from a series of corresponding mea-sured values of a random sample from a population on the basis of a specifi ed procedure. This is a purely statistically defi ned value which describes the distribu-tion of this substance in the environmental medium in question for a defi ned population at the time when the study was carried out. The reference value is to be specifi ed as a numerical value which is uniquely defi ned by the statistical pro-cedure applied and the population investigated at a particular point in time. The (upper) reference value is the 95th percentile of the concentration of the substance in the environmental medium investigated for the reference population. Defi ni-tion on the basis of the 95th percentile is an internationally accepted convention. An additional important item of information is the size of the random sample from which the reference value has been derived. The reliability of the statistically obtained 95th percentile can for example be described by providing the 0.95 con-fi dence interval of the 95th percentile.

9.4 Evaluation of Indoor Air Quality with the Aid of Guide Values

The aim of toxicologically derived guide values is to provide the user in scientifi c and offi cial practice with a numerical value which allows him to see from which concentration of the particular pollutant or group of pollutants in the indoor air a health risk to room users could, on the basis of current scientifi c fi ndings, no longer be excluded with suffi cient certainty or beneath which concentration any risk there might be could be ignored.

9.4.1 Requirements Relating to Guide Values for Indoor Air

9.4.1.1 Health Reference Guide values are derived on the basis of current scientifi c fi ndings about the health effects of the substance or group of substances under consideration. First of all the toxicological base data are obtained for non - carcinogenic substances and carcinogenic substances without an initiating effect (substances with non - stochastic effects) on the basis of appropriate animal studies and observations in humans. These are on the No Observed Adverse Effect Level ( NOAEL ) and Lowest Observed Adverse Effect Level ( LOAEL ). Depending on the scope and quality of the toxicological data, the LOAEL value is linked with safety margins and extrapo-

9.4 Evaluation of Indoor Air Quality with the Aid of Guide Values 193

lation factors in accordance with the Basic Scheme in order to take into account uncertainties on the one hand and specifi c physiological circumstances and risk groups (children, for example) on the other hand (Ad - hoc AG, 1996 ). These guide values (RW = Richtwert) for indoor air are obtained under the condition of con-tinuous and all - day use of an indoor room by sensitive groups of persons. In exceptional cases guide values have also been given for different periods of time (CO and NO 2 ) (see Table 9.1 ). In the case of indoor air concentrations above Guide Value II health effects on sensitive room users can no longer be excluded with suffi cient probability. This guide value RW II thus defi nes a danger threshold (Ad - hoc AG, 1996 ).

Table 9.1 German guide values ( RW I and RW II ) for indoor air quality (Ad - hoc AG ) a .

Substance RW II RW I Year of derivation

Toluene 3 mg/m 3 0.3 mg/m 3 1996

Dichloromethane 2 (24 h) mg/m 3 0.2 mg/m 3 1997

Carbonmonoxide 60 (1/2 h) mg/m 3 6 (1/2 h) mg/m 3 1997 15 (8 h) mg/m 3 1.5 (8 h) mg/m 3

Pentachlorophenol 1 μ g/m 3 0.1 μ g/m 3 1997

Nitrogen dioxide 0. 35 (1/2 h) mg/m 3 – 1998 0.06 (1 week) mg/m 3

Styrene 0.3 mg/m 3 0.03 mg/m 3 1998

Mercury (Hg - vapour) 0.35 μ g/m 3 0.035 μ g/m 3 1999

Diisocyanates b No recommendation of RW I or RW II possible 2000

Tris(2 - chloroethyl) - phosphate 0.05 mg/m 3 0.005 mg/m 3 2002

Bicyclic Terpenes ( α - Pinene) 2 mg/m 3 0.2 mg/m 3 2003

Naphthalene 0.02 mg/m 3 0.002 mg/m 3 2004

Aliphatic Hydrocarbons (C9 - C14) 2 mg/m 3 0.2 mg/m 3 2005

TEQ (dl - PCBs) 5 pg/m 3 2007

Chlorinated Biphenyl 118 0.01 μ g/m 3 2007

a Regularly updated at: http://www.umweltbundesamt.de/gesundheit/innenraumhygiene/richtwerte - irluft.htm. b Wolf and Stirn (2000) .


9.4.1.2 Legal Reference In general legal terms in Germany ‘ danger ’ is understood as a set of circumstances which if unimpeded can objectively be expected to develop within a foreseeable time and with suffi cient probability into damage being suffered by an object pro-tected under law. Here the more important the object protected under law (human life, for example) and the greater the damage to be expected (health effects, for example) the greater will be the requirements relating to the probability of the damage not occurring. Warding off a situation of this kind is referred to as danger prevention.

A distinction should be drawn between legally specifi ed limit values and guide values which are not legally binding. Guide values can however be matters for court decision either de facto or in execution when there is a reference to a legal regulation (e.g., in enforcement of building codes). Legal applicability (in admin-istrative proceedings, for example) is ensured by the fact that danger - related aspects are recognized (e.g., by including the LOAEL as a point of departure in obtaining the German RW II guide value) during the scientifi c derivation of guide values which has a reference to a particular legal framework.

9.4.2 Basic Scheme for Deriving Guide Values for Indoor Air

With the publication of the so - called basic scheme of the IRK/AOLG Ad - hoc working group, the German Federal Environment Agency and the Supreme State Health Authorities in Germany have provided a guideline for estimating the risks arising from contamination of indoor air (Ad - hoc AG, 1996 ; Seifert et al ., 1999 ).

The working group fi rst defi nes a level at which a health hazard would probably exist and by introducing safety margins a level where there would be no concern for health effects. Essential steps of the guide value derivation are (i) identifi cation of an adverse effect threshold (LOAEL) from human or animal studies, (ii) inter-species conversion from animal to human, (iii) consideration of sensitive individu-als (intra species factor), (iv) conversion from short term to continuous exposure, (v) physiologic differences within the population (i.e., children factor), and (vi) consideration of pathways other than inhalation. In Figure 9.1 standard default factors are given, which are used unless there is suffi cient information that justi-fi es another factor. Since 1996 guide values have been derived on this basis for a number of substances in the room air.

RW II is an effect - related value based on current toxicological and epidemiological knowledge of a substance ’ s effect threshold which takes uncertainty factors into account. It represents the concentration of a substance which, if reached or exceeded, requires immediate action as this concentration could pose a health hazard, especially for sensitive people who reside in these rooms over long periods of time. Depending on the mode of action, guide value II may be defi ned either as a short - term value (RW II K) or a long - term value (RW II L).


What is meant by action here is a need for testing without delay, for example, with regard to decisions to carry out remediation to reduce exposure. It may there-fore be necessary for rooms to be closed. If guide value RW II is exceeded this should immediately result in check measurements under normal utilization con-ditions and – if possible and appropriate – determination of the internal exposure of room users.

RW I represents the concentration of a substance in indoor air for which, when considered individually, there is no evidence at present that even lifelong exposure is expected to have any adverse health impacts. Values exceeding this are associ-ated with an abnormal exposure which is undesirable for health reasons. For precautionary reasons, there is also need for action in the concentration range between RW I and RW II. RW I is derived from RW II through the introduction of an additional factor (usually 10). This factor is a convention. In the case of strong - smelling substances, RW I must be obtained not on the basis of this sche-matic derivation but on the basis of odor perception (detection threshold) if this results in a lower numerical value. RW I can serve as a target value for clean - up operations. It should be undercut rather than merely complied with.

Guide values for individual substances derived using the basic scheme of the AG IRK/AOLG Ad - hoc working group provide no information regarding the pos-sible effects of combination with different substances. To date the guide values for individual substances and groups of substances that have been established by the Ad - hoc working group are set out in Table 9.1 :

The advantage of these guide values for indoor air is their uniform derivation with its orientation toward risk. Building construction law (building codes) repre-

Figure 9.1 Derivation of guide values according to the basic scheme of the Ad - hoc AG (1996) .


sents the major legal framework when defi ning guide values for room air. Accord-ing to state building regulations a building must be built and maintained in such a way that the user is not exposed to health risks or signifi cant nuisances. Should RW II be exceeded, dangers to health can by defi nition be expected, particularly for sensitive individuals such as pregnant women, babies and small children. Exceeding RW I points to an increased and, from the health point of view, undesir-able exposure.

9.4.3 Application of the Guide Values in Risk Management

Consequent upon RW I or II being exceeded, possible options include utilization recommendations and/or technical measures relating to the building itself. The aim of utilization recommendations and immediate action for a room is to prevent increased exposure of users to pollutant concentrations above RW II. The aim of long - term measures is for the air in the room in question to be brought down below RW I.

Under the assumption that the measurement results available are reliable, once it is noted that RW II has been exceeded utilization recommendations should be issued to avert possible health risks to sensitive individuals and action should be taken without delay to reduce exposure. If the source of the contamination has been identifi ed, it should as a rule be removed, which may mean building remodel-ing if the source is found to be in the building itself. If the source cannot be removed in the short term, then, depending on local conditions and after consulta-tion with the relevant health authorities, temporary or alternative measures are indicated (sealing, partitioning off, painting and/or restrictions or prohibition in using the area). In this case the location of the source(s) and the action taken must be clearly documented.

If RW I is exceeded (in the range between RW I and RW II) an immediate endangerment of the health is not to be expected. However people may report feelings of malaise, particularly in the case of perception of odors, and health impairments may occur. According to current scientifi c thinking (Winneke, 1994 ) these impairments do not result in adverse effects on the health, but under certain circumstances such as a repeated or long - lasting infl uence may represent an unac-ceptable nuisance.

From the point of view of a proportionate response, building or other source - related changes should not be the fi rst step of action in the concentration range between RW I and RW II but instead greater ventilation should primarily be secured and depending on the individual case (with SVOCs, for example) more intensive cleaning carried out. In the case of dustborne semi - volatile substances, exposure to fl oor dust should be reduced by regular dust - bonding cleaning (e.g., wiping with a damp cloth). If, despite documentably more intensive ventilation, check measurements after a certain time (usually after one month) indicate no detectable improvement in the air quality and RW I continues to be exceeded, the next step, which also applies to concentrations in the range between RW I and RW II, would be to take further measures, which could even include technical


building alterations since higher exposure extending over a relatively long period of time (more than 12 months) is not acceptable for precautionary reasons.

By defi nition the danger - related RW II guide values for indoor air are derived with the condition of a continuous and all - day utilization of an indoor room by sensitive groups of individuals. Within the context of risk management (with the aim of preventing danger), it can be advisable to take into account the duration of exposure in specifi c indoor rooms when interpreting RW II. For the overwhelming majority of rooms it is not permissible to convert the guide value concentration taking into consideration an exposure time which is reduced in comparison with the full 24 hours. Taking it into consideration does however become a possibility for rooms whose special function means they are only used for a limited period of time each day. Conversion of the guide value is only possible when this is permitted by knowledge of the mechanism behind the health effects caused by the substance in question. It must however also be borne in mind that for particular indoor environments such as sports halls, for example, the higher levels of physical activity there mean increased respiratory minute volumes. The procedure in individual cases must be made adequately transparent as part of risk communication.

9.4.4 Recommendation

To decide whether a guide value has been exceeded or fallen short of, a check measurement should, as specifi ed in the basic scheme, be taken under utilization conditions before initiating further action.

9.4.5 Guide Values by the A d - hoc WG Not Based on RW I and RW II

For over 150 years, carbon dioxide has been an acknowledged indicator of indoor air quality. To estimate the air quality in mechanically ventilated buildings, DIN EN 13779 (DIN, 2007 ) proposes four different levels of indoor carbon dioxide concentration. However, apart from the early guideline value of 1000 ppm carbon dioxide recommended by Pettenkofer in 1858, there is no actual guideline value for naturally ventilated buildings. Regarding recent intervention studies, the Ad - hoc AG therefore recommends the following guide values, based on health and hygiene considerations: concentrations of indoor air carbon dioxide below 1000 ppm are regarded as harmless, those between 1000 and 2000 ppm as elevated and those above 2000 ppm as unacceptable. In addition to the recommendations for TVOC values, this further assists in the assessment of indoor air quality (Ad - hoc AG, 2008a ).

When evaluating the health effects of indoor air fi ne particulate matter, the indoor dynamics as well as the physical, chemical and biological properties of fi ne particles have to be considered. The indoor air fraction PM 2.5 largely stems from outdoor air. Accordingly, the Ad - hoc AG also recommends WHO ’ s (2006) 24 - hour mean guideline value of 25 μ g/m 3 PM 2.5 for indoor air evaluation. In contrast to PM 2.5 , coarse particles (PM 10 ) in schools, kindergartens and dwellings show much


higher indoor air concentrations. Additional sources indoors have to be assumed. Because of the different composition of indoor air compared to outdoor air and due to the lack of dose - response relationships of coarse particles in indoor air, the health effects of indoor air PM 10 can not be evaluated yet. Suffi cient and consistent ventilation is an indispensable basis to reduce PM concentrations in indoor spaces. Furthermore, known sources of PM indoors should be detected consequently and subsequently minimized (Ad - hoc AG, 2008b ).

9.5 Health Evaluation with the Aid of the TVOC Concept

Indoor air always contains a large number of different substances. According to the evaluation dealt with in Section 9.4 which is based on guide values for indi-vidual substances, this situation can be assessed with the aid of the total concentra-tion of volatile organic compounds (TVOC) (Seifert, 1999 ). As early as the beginning of the 1980s attempts were also made to determine the health signifi cance of VOC mixtures (M ø lhave, Bach and Pederson, 1986 ). Experimental investigations with specifi ed mixtures of solvents and also practical experience showed that as the TVOC concentration rises, so too does the probability of reactions and negative effects on the health (ECA, 1997 ). The question of evaluating the health risk and also of evaluating a possible odor burden could not however be answered since this did not depend on the totality of the VOCs but decisively on the type and proportions of individual compounds.

The composition of the mixture of substances to which the measured variable TVOC refers differs considerably from case to case (M ø lhave et al ., 1997 ). Due to the different composition of the mixture of substances found in indoor air the TVOC value can have no concrete toxicological basis. It can be used for character-ization of exposure and when searching for sources and also in risk assessment as a screening parameter for a possible sensory irritation (M ø lhave, 2003 ). The TVOC concept (Seifert, 1990, 1999 ) was based in principle on the statistical evalu-ation of the data from a previous environmental survey in Germany in which the air in living rooms was investigated (Krause, Chutsch and Henke, 1991 ). TVOC value can thus be interpreted in the sense of reference values since in studies TVOC concentrations around 0.3 mg/m 3 fall within the 50th percentile and around 1 mg/m 3 approximates a 95th percentile. An uncritical use of the TVOC values may result in erroneous conclusions being drawn (Oppl, H ö der and Lange, 2000 ).

9.5.1 Recommendation Relating to the Application of TVOC Values

The following recommendation regarding the application of TVOC values renders Seifert ’ s TVOC concept of 1999 in more precise terms (Seifert, 1999 ). The scheme recommended in Table 9.2 is subdivided into fi ve levels which allow room for interpretation in their transitional regions (Ad - hoc AG, 2007 ). These levels will be

9.5 Health Evaluation with the Aid of the TVOC Concept 199

Table 9.2 Recommendations relating to the application of TVOC values ( A d - hoc AG , 2007 ).

Level Concentration range (mg/m 3 )

Health evaluation Questions for clarifi cation

Recommendations

1 ≤ 0.3 mg/m 3 No hygienic objections and no health - related concerns. Usually no complaints/symptoms.

Guide values being exceeded?

No further action

2 > 0.3 – 1 mg/m 3 Still no relevant health - related concerns, provided no guide values for individual substances or substances groups are exceeded. In individual cases, complaints/symptoms or odor perceptions, for example, after minor renovation work or installation in recent weeks of new furniture

Guide values being exceeded? Reference values being noticeably exceeded? Are there any objections to indoor climatic conditions (air exchange, temperature, air humidity)?

Ventilate adequately especially after renovation work. Determine VOC sources (examine the room, for example). Check the use of cleaning agents. Re - measurements to check the guide value violation under utilization conditions.

3 > 1 – 3 mg/m 3 Some objections and distinct health issues. Utilization of rooms which are regularly used is only acceptable to a limited extent ( < 12 months). Within 6 months or so, TVOC concentrations should have fallen clearly below the TVOC value initially measured. Cases with complaints/symptoms or odor perceptions, for example, after major renovation work.


Check guide value violations immediately by means of re - measurements under utilization conditions and take into consideration the information in Section 9.4 . Check the relevance to health of instances of reference values being noticeably exceeded. In all cases search for sources and check the ventilation situation: ventilate intensively and if necessary lay down utilization and ventilation conditions. Re - measurements after 1 month or so are recommended (under utilization conditions). If, despite the efforts described, the TVOC concentration after 12 months is still above 1 mg/m3, adequate remodeling will need to be included in future planning.




Recommendations

4 > 3 – 10 mg/m 3 Major objections and health concerns. Utilization of rooms which are regularly used is only acceptable to a limited extent ( < 1 month). Within one month the TVOC concentration should have fallen below 3 mg/m 3 . Cases with accumulation of complaints/symptoms or odor perceptions, for example, after major renovation work.


Check guide value violations immediately by means of re - measurements under utilization conditions and take into consideration the information in Section 9.4 . Check the relevance to health of instances of reference values being noticeably exceeded. A toxicological evaluation of individual substances or groups of substances is necessary. In all cases search for sources and ventilate intensively; and if necessary lay down utilization and ventilation conditions and initiate suitable minimization measures. Any occupancy of the room which may be necessary can only be tolerated with a time restriction for each day over a maximum period to be specifi ed by the health authorities (limited hours per day and time period). Re - measurements after 1 month or so are recommended (under utilization conditions). If, despite the efforts described, the TVOC concentration after one month is still above 3 mg/m 3 , adequate remodeling measures will need to be included in future planning.

Table 9.2 Continued

9.5 Health Evaluation with the Aid of the TVOC Concept 201



Recommendations

5 > 10 mg/m 3 Not acceptable, serious health concerns. As far as possible avoid room utilization. Occupancy is only permissible with time restrictions in hours per day and over a specifi c time period. When values exceed 25 mg/m 3 any use of the room should be prohibited. Within one month the TVOC concentration should have fallen below 3 mg/m 3 . Usually complaints/symptoms or odor perceptions, for example, after misapplications, accidents.

Guide values being exceeded? Reference values being exceeded? Are there any objections to indoor climatic conditions (air exchange, temperature, air humidity)?

Check guide value violations immediately by means of re - measurements under utilization conditions and take into consideration the information in Section 9.4 . Check the relevance to health of instances of reference values being noticeably exceeded. A toxicological evaluation of individual substances or groups of substances is necessary. In all cases search for sources and ventilate intensively; and lay down utilization and ventilation conditions and initiate suitable minimization measures. Any occupancy of the room which may be necessary can only be tolerated with a time restriction for each day over a maximum period to be specifi ed by the health authorities (limited hours per day and time period). Re - measurements within 1 month or so are recommended (under utilization conditions). If minimization measures bring the concentration below 10 mg/m 3 in the period under consideration but it still remains above 3 mg/m 3 , the recommendations for action given under Level 4 will apply. If, despite the efforts described, the TVOC concentration after one month is still above 10 mg/m 3 , room use should be prohibited and adequate remodeling measures initiated.

Table 9.2 Continued


briefl y introduced below and shown in the table with further information and recommendations. The recommendations should be taken as cumulative – in other words, the recommendations of one level will also apply – where applicable – in the next higher level as well.

One precondition of applying the scheme is that the toxicologically based guide values of individual substances are not thereby exceeded. A separate evaluation will however always be required when substances with low odor perception thresh-olds are involved which even at lower concentrations can be bothersome due to their odor activity or when noticeably high concentrations of individual substances occur.

Level 1: There are no health objections to TVOC values below 0.3 mg/m 3 provided no guide values are exceeded. They are referred to as ‘ target values ’ (health precau-tions range) and can be reached once a suffi cient period of time has elapsed fol-lowing new building construction or renovation work in rooms – or they can, if possible, be undershot.

Level 2: TVOC values between > 0.3 and 1 mg/m 3 may still be classifi ed as being free of health objections provided no guide values are exceeded. This concentration range points, for example, to emitted solvents not being fully removed by ventila-tion and indicates the necessity of an increase in ventilation.

Level 3: TVOC values between > 1 and 3 mg/m 3 are to be evaluated as warning signs as far as health is concerned and will apply for a limited period of time ( < 12 months) as the top limit for rooms which are intended to be occupied for relatively long periods of time. A TVOC concentration under utilization conditions of 1 mg/m 3 should not be exceeded for extended periods of time in normally used living rooms, school rooms or offi ces unless there has been recent renovation work or new fur-niture has been installed. Values of this kind would be interpreted as an indication of an additional and possibly unwanted entry of VOCs. The relevance to health of instances of reference values being noticeably exceeded should be checked. An individual toxicological evaluation of at least those substances with the highest concentrations is recommended. Subsequent re - measurement to check the indoor air quality is carried out under utilization conditions.

Level 4: Rooms with TVOC values between > 3 and 10 mg/m 3 are evaluated as giving cause for concern regarding health and should, where no alternatives are available, only be used for a limited period of time (no more than 1 month) and with a higher frequency of regular ventilation. A toxicological evaluation of individual substances and groups of substances should be carried out. Subsequent re - measurement to check the indoor air quality is carried out under utilization conditions.

Level 5: TVOC values between > 10 and 25 mg/m 3 are classed as unacceptable from the health point of view. Use of the room is to be avoided as a rule but daily occupancy on a temporary basis (less than 1 hour per day) is, if need be, permis-sible and provided there is a higher frequency of regular ventilation. At values

9.6 Evaluation of Indoor Air Quality with the Aid of Reference Values 203

> 25 mg/m 3 no use of the room is to be allowed. Subsequent re - measurement to check the indoor air quality is carried out under utilization conditions.

9.5.2 Time Curve of Higher TVOC Concentrations

Evaluation with the aid of reference values or the TVOC value is particularly suit-able for rooms in which an equilibrium concentration has established itself, and when high levels or patterns of individual compounds occur may even allow con-clusions to be drawn regarding particular sources. If sources have recently been brought into a room, such as is the case after renovation work, new construction, remodeling activities and the installation of new furnishings, indoor air concentra-tions will in most cases be high initially and then with time decay more or less rapidly to an equilibrium level. For many substances concentrations can even rise after a certain period of time. The period of time between when measurements were taken and when the activity was carried out (renovation, for example) will therefore have a signifi cant effect on the result. Experience shows that an equilib-rium level of the concentration may take months, even as long as a year or so, to become established. For this reason higher TVOC concentrations can be tolerated for a limited time (no more than 12 months) in new buildings, freshly renovated rooms or following the installation of new furniture or furnishings. However, values of 3 mg/m 3 should not be exceeded (see Table 9.2 ). In this connection special ventilation recommendations for the utilization should be formulated. Over a period of 6 months following the construction of new buildings or renova-tion, the higher TVOC values, which were measured under standardized condi-tions, should have fallen markedly. The time curve for concentration can be documented by means of repeated measurements and a time curve forecast drawn up.

Ventilation must always form part of the normal utilization of rooms which people use. Particular circumstances, such as following renovation, may make increased ventilation activity necessary and this may need to be agreed with the room users.

9.6 Evaluation of Indoor Air Quality with the Aid of Reference Values

With a comparative evaluation the measured value obtained by a measurement is compared with information already collected. In the most favorable case, current frequency distributions as well as characteristic parameters (reference values) derived statistically in accordance with specifi c standards will be available which describe the normal occurrence of a substance in an environmental medium. Since it is defi ned exclusively statistically, a reference value is not linked with any health - related evaluation. If, for example, the concentration lies below the reference value, this does not imply any health - related evaluation but rather only that the vast major-ity of the population is exposed to a comparable order of magnitude.


In this connection a methodological paradox is inherent in the system. Since reference values as (dynamic) statistical quantities represent the presence of sources and the daily use of substances and products in the indoor environment for the reference time period, reference values are primarily applicable to sub-stances which have had a wide application for many years. In contrast a compa-rably broad database is not available for substances and substitute products which have only recently become more frequently used. If reference values for this kind of recently introduced compound are derived from existing (historical) data records, they will often only be a little higher than the detection limits and the use of a ‘ new ’ substance in the indoor environment will necessarily mean the reference value being exceeded. The validity of a reference value derived in this way is low and the hidden danger of using values of this kind is that sensible substitute substances cannot be introduced and the development of innovative products is impeded. It follows from what has been said that reference values can change over the course of time. In recent decades marked changes have been observed even in the range of substances occurring in the indoor environment and in their con-centrations in the room air and in the house dust. For this reason reference values should be updated as close to real time as possible in the fi eld of indoor measurements.

9.6.1 The Current State of Indoor Air Reference Values

The fi rst representative reference values for indoor air in German households were derived from a study conducted by the old Federal Health Offi ce into around 500 households in West Germany in 1985/86 (Krause, Chutsch and Henke, 1991 ). In 2007 updated reference values for individual volatile organic compounds (VOCs) and for aldehydes ( Tables 9.3 and 9.4 ) were derived from long - term sampling in > 550 households by means of diffusion samplers. Additionally a number of mostly regional studies have been made in schools (Heinzow et al ., 1994 ), day - care centers (Schreiner, Wetzel and Kirchbach, 2001 ) and offi ces (Schlechter et al ., 2004 ).

In contrast, some contractors and appraisers use values derived from their own data (mainly from causes for a complaint), in some cases with different designa-tions (Scholz, 1998 ; Schleibinger et al ., 2002 ; Hippelein, 2003 ; AG Ö F, 2007 ). Data from VOC - indoor - air measurements of the past fi ve years were collected by Hofmann and Plieninger (2008) and entered into a database. 2663 datasets with 300 129 single measurements from the years 2002 to 2006 have been evaluated. Statistical values as means, medians and certain percentiles are stated for more than 300 substances.

9.6.2 Recommendations

If a reference value is exceeded the fi rst step should be to carry out a plausibility check. This check will cover the measurement and ambient conditions, sampling,

Table 9.3 Selected reference values ( μ g/m 3 ) for VOC s in indoor air in German homes (n = 555) ( UBA , 2008 ).

VOC n < LOD a Reference value b Median c Maximum

n - Hexane 204 22.8 1.4 414 n - Octane 257 10.3 1.1 69.4 n - Nonane 293 12.1 < 1 66.4 n - Decane 222 14.9 1.4 108 n - Undecane 220 14.8 1.4 135 n - Dodecane 255 7.9 1.1 186 n - Tridecane 323 4.2 < 1 44.0 n - Tetradecane 86 5.4 1.7 55.4 n - Pentadecane 204 3.7 1.2 21.1 n - Hexadecane 298 2.3 < 1 9.7 Cyclohexane 152 39.1 2.5 456 Methyl - cyclohexane 266 26.5 1.1 400 Benzene 137 7.7 1.8 61 Toluene 0 57.6 13.5 2400 Ethyl - benzene 179 6.8 1.4 40.8 m - , o - , p - Xylene 52 21.2 4.5 248 Iso - + n - Propyl - benzene 454 2.6 1.4 7.4 2 - Ethyl - toluene 475 2.3 < 1 13.6 3 - + 4 - Ethyl - toluene 341 8.3 1.4 60.9 1,2,3 Trimethyl - benzene 446 2.9 < 1 12.3 1,2,4 - Ethyl - benzene 177 10.3 1.5 58.8 1,3,5 - Ethyl - benzene 449 2.9 < 1 24.3 Styrene 329 4.8 < 1 32 Naphthalene 517 1.2 < 1 4.9 Ethylacetate 12 70.8 9.3 Butylacetate 64 30.7 4.1 1 - Methoxy - 2 - propanolacetate 426 3.6 < 1 Methylethylketone 519 9.2 < 7.5 Methylisobutylketone 407 2.6 < 1 1 - Butanol 10 17.6 5.4 Isobutanol 505 4.9 < 3.5 2 - Methoxyethanol 528 1.2 < 1 2 - Ethoxyethanol 502 1.5 < 1 2 Butoxyethanol 215 10.3 1.4 2 - Butoxyethoxyethanol 374 6 < 1 2 - Phenoxyethanol 357 3.7 < 1 1 - Methoxy - 2 - propanol 198 8.4 1.5 1 - Butoxy - 2 - propanol 295 12.8 < 1 1 - Phenoxy - 2 - propanol 544 < 1 < 1 2 - Ethyl - 1 - hexanol 86 11.4 2.6 Dipropyleneglycol - monobutylether 457 1.6 < 1 Texanol 407 2.8 < 1 TXIB 227 5.5 1.2 α - Pinene 8 67.6 9.8 800 β - Pinene 240 8.3 1.2 47.8 Limonene 38 103 11.5 400 3 - Carene 141 22.7 2.6 336 Longifolene 465 1.8 < 1 8 TVOC mg/m 3 555 1.1 3.0 0.3

a LOD = Limit of Detection. b 95th percentile. c 50th percentile.


Table 9.4 Reference values ( μ g/m 3 ) for Aldehydes in indoor air in German homes (n = 586) ( UBA , 2008 ).

Aldehyde n < LOD a Reference value b Median c Maximum

Formaldehyde 0 47.7 23.5 68.9 Acetaldehyde 8 50.3 15.5 863 Propanal 0 6.1 2.5 40.9 Butanal 4 8.1 2.4 43.2 Pentanal 7 10.6 3.7 28.4 Hexanal 0 30 9.8 91.7 Heptanal 68 3 1.3 9.1 Octanal 3 3.6 1.6 6.8 Nonanal 0 14.7 7.2 30.2 Decanal 17 5.5 2.5 9.9 Undecanal 327 3.1 < 1 13.2 Furfural 21 2.8 0.9 19.3 Benzaldehyde 13 6.6 2.9 11.9 Isovaleraldehyde 384 3.9 < 0.5 15.9 Methylglyoxal 295 17.8 < 1 54.1

a LOD = Limit of Detection. b 95th percentile. c 50th percentile.

the quality of measurement (quality assurance) and whether it is plausible for the substance to be found in the indoor air if, for example, identifi cation of the sub-stance had only been on the basis of a suggestion from a mass - spectrum library. Taking into consideration the level of the concentration found and the toxicological properties of the individual substance in question, it should be determined whether exceeding the reference value has any relevance to health so that check measure-ments can be taken if necessary.

If it is confi rmed that the reference value has been exceeded and this is assessed as having relevance to health, it is recommended that matters be taken further with a detailed investigation which should include a search for causes/sources, a concentration/time curve forecast and if necessary an individual toxicological analysis in order to examine and decide whether further action should be taken.

9.7 Application of Measured Values in Order to Evaluate Indoor Air Quality

The degree to which the public health service acts successfully depends on accep-tance of offi cial procedures. This requires transparency in presenting the risk assessment undertaken and the risk management provided. A transparent

9.8 Evaluation of Substances Without Reference Values 207

approach to the health - related evaluation of contaminants in indoor air will pri-marily succeed when a systematic procedure is in place. As a rule this includes three main steps. Following an evaluation of the quality and documentation of the measurement which has been carried out, particularly with regard to the measure-ment strategy, boundary conditions, measurement methodology and quality assur-ance a check is made, on the basis of guide values and the TVOC concept: this is followed by a comparative evaluation of the measurement results in the light of reference values. If a case of a guide value being exceeded is found or an exposure which is greater than normal, the next step is to carry out a health risk assessment.

The only laboratories which should be commissioned to carry out measure-ments are those which have a documented quality system in place, which can supply evidential documentation regarding the procedure used in the identifi ca-tion of the VOC and the type and frequency of calibration, and which participate successfully in external inter laboratory tests and/or comparative laboratory inves-tigations (cf also the corresponding section on ‘ Quality assurance ’ in the various parts of the VDI 4300 series of guidelines, VDI 4300, 1997 ). Attention is also drawn to the contents of standard DIN EN ISO/IEC 17025 (2005) which deals with general requirements applicable to testing and calibration laboratories and to the Guide to the Expression of Uncertainty in Measurement ( GUM ) published in 1995 by the International Organization for Standardization ( ISO ) (see also ‘ Guide to the Expression of Uncertainty in Measurement ’ , DIN preliminary standard ENV 13005, 2nd edn: 1999 - 06) (ISO, 1995 ). A guide to documenting the information to be recorded during indoor air investigations is provided as Appendix D to DIN EN ISO 16000 - 1 (2006) .

9.8 Evaluation of Substances Without Reference Values From the IRK / AOLG A d - hoc Working Group

A large number of guide values from other institutions and individual authors is in existence, they also vary considerably in their scientifi c derivation and legal signifi cance. An overview of individual values, the underlying studies and uncer-tainty factors and also references to sources will be found on the TERA website (Toxicology Excellence for Risk Assessment under ITER International Toxicity Estimates for Risk Database at http://www.tera.org/iter/ ).

In addition to the hazardous - substance - related limit values in TRGS 900 (TRGS 900, 2000 ) which apply to the workplace area, for a limited number of substances toxicologically derived guide values for evaluating air concentrations are also avail-able such as those of, for example, the WHO and other organizations. Here a distinction is not always drawn between outdoor air (ambient air) and indoor air. The values from these other organizations, including a monograph (Calabrese and Kenyon, 1991 ), are based on published toxicological derivations. They contain the


requisite basic data and have the required transparency which means that a pro-visional derivation using the basic scheme (see Section 9.4.2 ) may if required be undertaken in accordance with the aforementioned criteria of German building construction law. The values themselves should not be used without considering and checking the suitability of the individual foundations and derivations, includ-ing the safety factors selected for evaluating indoor air exposures.

These include, for example:

1. Air Quality Guidelines of the WHO (WHO, 2000 ) http://www.who.int/air/en/

2. Recommendations of the EU project INDEX (Kotzias et al ., 2005 ) http://www.jrc.cec.eu.int/more_information/download/indexproject.pdf

3. Minimal risk level (inhalation) of the ATSDR (ATSDR, 2006 ) http://www.atsdr.cdc.gov/glossary.html

4. Reference concentration (RfC) of the US EPA (2006) http://www.epa.gov/iris/gloss8.htm

5. Risk exposure level s ( REL ) of the Californian OEHHA (OEHHA, 2003 ) http://www.oehha.ca.gov/air/chronic_rels/pdf/relsP32k.pdf

6. Tolerable concentration ( TC ) of Health Canada (Health Canada, 2006 ) http://www.hc-sc.gc.ca/ewh-semt/pubs/contaminants/index_e.html

7. Tolerable concentration in air ( TCA ) of the RIVM (Baars et al ., 2001 );

8. Effect - related indoor guide values (WIR) of the Austrian Ministry of the Environment and the Austrian Academy of Sciences (BMLFUW, 2006 ) http://www.innenraumanalytik.at/fr_texte.html

As regards carcinogenic substances, risk value based on the concentration unit of 1 μ g/m 3 (so - called ‘ unit risk ’ or UR) are available. Due to the general lack of social acceptance regarding the level of risk, URs can in practice only be applied to a limited extent. For substances without a detectable effect threshold the minimiza-tion rule will usually apply.

Acknowledgment

The chapter is based on recommendations prepared by the Ad - hoc Working Group published in 2007 (Ad - hoc AG, 2007 ). We wish to express our gratitude to the members of the group and the secretariat at the Federal Environment Agency (UBA).

References 209

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Field Studies


Part Three

Effect of Ventilation on VOC s in Indoor Air Kwok Wai Tham , S. Chandra Sekhar , Mohamed Sultan Zuraimi

10

10.1 Introduction

Volatile organic compound s ( VOC s) and carbon dioxide (CO 2 ) are two major indi-cators of indoor air pollution in air - conditioned offi ce buildings. Indoor air quality assessments would include VOCs and CO 2 measurements and attempt to identify their sources so that strategies for effective control may be implemented. Options for control may be source control and ventilation. Source control is the more effec-tive while ventilation is potentially expensive, given the unpredictable (and usually rising) cost of energy.

In some of the studies conducted, the effi cacy of ventilation to control VOC levels has been noted to vary widely depending on the operation and maintenance of the building, the presence of sources and their source strengths and other physical and chemical processes determining the VOC behavior (Hodgson et al ., 2003 ; Sax et al ., 2004 ; Zhang et al ., 2007 ; Jia, Batterman and Godwin, 2008 ). These studies suggest that source control measures, in addition to adequate ventilation, are required to limit concentrations of VOCs in offi ce buildings. Conditioning of ventilation air for tropical buildings consumes considerable energy (Sekhar et al ., 2002 ) where recent construction practices have attempted to minimize the amount of conditioned air used for ventilation while still maintaining adequate indoor air quality. This chapter focuses on the relation between ventilation and VOCs in the hot and humid tropical climate.

This chapter presents a series of case studies to illustrate the VOC profi les and CO 2 concentration levels in a tropical climate, Singapore. These studies have been conducted over a period of fi ve years with buildings that use different ventilation systems. Emission rates of alkanes, aromatics, alcohols, cyclic alkanes, carbonyls and other VOCs were determined, and compared with similar studies conducted in Europe and North America. The sensory and health effects of these levels of VOCs are briefl y described. Cluster analyses were performed yielding plausible sources which are human or building material and building operation related.


215

216 10 Effect of Ventilation on VOCs in Indoor Air

Apportionment of VOCs source strengths to these sources are obtained through a mass balanced model.

Ventilation is often adopted as a control measure to keep VOCs and CO 2 below specifi ed threshold levels. The effect of operation of the ventilation system on spatial and temporal profi les is illustrated from two separate buildings. With adequate ventilation, the CO 2 levels are easily controlled within threshold levels. However, VOCs present a problem due to their build - up overnight between the time of ventilation shutdown after offi ce hours and its reactivation in the morning just before occupancy. Purging the building as an effective means to reduce the exposure risks is discussed and illustrated. Some concluding remarks on other changes to indoor exposures are made at the end of the chapter.

10.1.1 Building and Ventilation Characteristics of Offi ce Buildings in a Tropical Climate

The year - round warm and humid climate, combined with the pressure of space in a city within a knowledge - based and modern economy, has resulted in offi ce buildings which have light facades, large fenestrations with a reliance on air - conditioning to provide a congenial indoor climate for work. The high enthalpy of outdoor air makes it uneconomical and unfeasible to introduce a large amount of outdoor air, especially as pollution sources and activities are increasingly being introduced. Moreover, the volatility of oil prices, which affect the energy costs of air - conditioning, results in ventilation being used to satisfy building regulatory requirements. High recirculation of indoor air - conditioned air, with minimal outdoor ventilation, has been the commonly adopted strategy. Only recently has research been undertaken on a systematic way to provide baseline information of the indoor air quality and ventilation characteristics of these buildings (Tham et al ., 1998 ; Sekhar et al ., 2002 ; Sekhar, Tham and Cheong, 2003b ).

From the baseline information, CO 2 , VOCs and ventilation rate profi les have been identifi ed. Although the levels do not generally exceed local threshold levels (ENV, 1996 ), it is important to understand how these profi les come to be, what their potential effects on sensory perception and health may be, what emission strengths they have and how they may be apportioned to various sources such as humans and human - related activities, building materials and the ventilation system.

10.2 VOC Concentration Levels in Eight S ingapore Buildings

This section discusses the concentration levels of individual VOCs found in typical tropical offi ce buildings, their potential health effects and possible sources. Eight buildings, considering their age, location and air distribution system were selected for the purpose of detailed characterization of VOCs. At the outset, the choice of selecting a ‘ measurement fl oor ’ on the basis of its association with a dedicated air handling unit ( AHU ) was emphasized. A ‘ measurement fl oor ’ was chosen such

10.2 VOC Concentration Levels in Eight Singapore Buildings 217

that an AHU and the entire occupied zones that it serves are included. In a ‘ mea-surement fl oor ’ , typically spanning about 400 – 800 m 2 in Singapore, 2 – 5 sampling points at the occupant breathing level, one at return air and one ambient point were chosen for sampling. All the measurements were carried out during occupied hours, when the AHU was in operation. This facilitated an evaluation of the ven-tilation characteristics resulting in an effective exposure control strategy of the indoor VOCs. In order to obtain a reasonable representation of a building three levels, at top middle and bottom were selected. Table 10.1 is a summary of the buildings investigated.

Active sampling using dual multisorbent (Carbopack C, Carbopack B and Carbosieve SIII) tubes over a period of eight hours was employed. The sampled VOCs were desorbed using an automated TD system and transferred via a heated fused silica line into a GC. The VOCs were separated on a low bleed capillary column, identifi ed and quantifi ed by their retention times and target and three qualifi er ions using mass spectrometric procedures. The details of the analytical procedures can be found elsewhere (Zuraimi et al ., 2006 ).

10.2.1 Concentrations

Although more than 100 compounds have been identifi ed and quantifi ed, a select group of 23 compounds were focused on the basis of the following criteria (i) their prevalence in Singapore commercial offi ce buildings (Sekhar et al ., 2001 ), (ii) adverse health effects arising from exposure in particular, toxicity, carcinogenicity

Table 10.1 Building characteristics.

Attributes

Location (number) 7 – suburban 1 – downtown

Smoking indoors Not allowed Mean age (years) 3.4 Mean fl oor area (m 2 ) 9432 Mean No. of fl oors 8.4 Mean occupancy 255 Surface material Ceiling (%) 100 – acoustic tiles Walls (%) 38 – wall paper

63 – painted wall Floor (%) 25 – linoleum

75 – carpet Ventilation system Principle (%) 100 – mixing Mean Recirculation 90%


Table 10.2 Concentrations of VOC s .

Compound Concentration ( μ gm − 3 )

Mean Standard deviation Range

Alkanes 2 - methypentane 6.4 5.0 1.5 – 17.0 3 - methylpentane 5.7 5.6 1.1 – 15.7 n - hexane 34.6 43.1 4.4 – 114.5 n - heptane 14.7 22.3 2.4 – 68.0 n - decane 24.8 28.6 1.3 – 85.7 n - undecane 32.9 41.7 2.7 – 114.8 n - dodecane 29.8 44.1 1.9 – 134.8 n - tetradecane 38.6 31.9 3.9 – 97.9

Aromatics Benzene 87.1 44.3 32.4 – 167.5 Toluene 287.3 518.3 112.9 – 1641.8 o - xylene 43.4 30.2 15.4 – 103.5 m / p - xylene 143.0 141.3 53.7 – 404.6 Naphthalene 143.9 93.0 42.9 – 310.0

Alcohols Ethanol 17.2 15.1 6.5 – 27.8 1 - butanol 4.4 2.6 1.4 – 6.3 2 - ethyl 1 - hexanol 30.6 40.0 1.0 – 244.4

Cyclic Alkanes Methylcyclopentane 8.1 14.5 0.1 – 33.9 Methylcyclohexane 36.4 34.7 2.9 – 98.8

Carbonyls 2 - propanone 16.5 26.3 0.1 – 63.1 Benzaldehyde 29.0 26.8 2.1 – 83.5

Others Isoprene 10.5 5.6 3.6 – 21.6 1,1,1 - trichlorethane 36.1 35.6 3.6 – 112.2 Limonene 65.1 52.9 4.7 – 144.7

and mutagenicity and (iii) the utility for tracer of possible sources. Table 10.2 reports the individual concentration of VOCs found in the different buildings.

Generally, aromatic compounds have high concentrations in Singapore build-ings. Indeed, this was confi rmed in another study where the aromatic levels in Singapore buildings were higher than those in European ones (Zuraimi et al ., 2006 ). The mean concentrations of toluene, m / p - xylene and naphthalene in Singapore buildings are typically an order higher in magnitude than other classes

10.2 VOC Concentration Levels in Eight Singapore Buildings 219

Table 10.3 Most frequently occurring and abundant VOC s .

VOCs most frequently occurring VOCs most abundant

Toluene m / p - Xylene Benzene o - Xylene o - Xylene 1,4 - Dichlorobenzene 1,1,1 - Trichloroethane 1,3,5 - Trimethyl benzene Styrene Toluene 1,4 - Dichlorobenzene Ethylbenzene Ethylbenzene Styrene 1,3,5 - Trimethyl benzene 1,1,1 Trichloroethane Cyclohexane Naphthalene m / p - Ethyl toluene n - Hexadecane Benzaldehyde 1,2,3 - Trimethyl benzene Isoprene Limonene Naphthalene Dibutyl phthalate Limonene Benzene Acetophenone Cyclohexane Dichlorotetrafl uroethane m / p - Ethyl toluene m / p - Xylene

of VOCs. Their values range from about fi fty to several hundred μ gm − 3 . In Singa-pore buildings, the mean concentrations of benzene, toluene, m / p - xylene, naph-thalene and benzaldehyde are 87, 288, 143, 144 and 29 μ gm − 3 respectively. From the more than 100 identifi able compounds, a list of ‘ most frequently found com-pounds ’ is prepared by adding up frequencies of every single compound from the main database (on top of the 23 compounds) and sorting them in descending order considering all the buildings (Sekhar et al ., 2001 ). Similarly, a list of ‘ most abun-dant compounds ’ is also presented in Table 10.3 . It is observed that the dominant class of VOCs in terms of occurrence and abundance in Singapore buildings are the aromatics.

The presence of toluene in most studies is in line with the fi ndings of other studies such as European audit and US review (Table 10.4 ). However, its presence toward the top of the list together with benzene may be due to the urban nature of Singapore. Figure 10.1 shows the benzene and toluene concentration measured for Singapore buildings in comparison with other countries. Concentrations of toluene in the three Nordic buildings, Denmark, Norway and Finland are 10, 62 and 7 μ gm − 3 respectively, while in continental Europe, the means of toluene con-centrations are 49, 12 and 27 μ gm − 3 for buildings in France, Germany and Switzerland respectively. Mean toluene concentrations in buildings from the UK and Greece are 69 and 47 μ gm − 3 respectively. The lowest mean toluene concentra-tion determined among the EU countries is from Finland. The mean toluene concentration in Singapore buildings has the highest value at 287 μ gm − 3 . It was found that the difference in toluene concentration in buildings in EU and Singa-pore was signifi cant.


Table 10.4 Comparison of ubiquitous indoor VOC s of Europe, North America and Singapore.

Singapore a European audit b US review c

1 o - Xylene Acetone o - Xylene 2 Toluene Isoprene Benzene 3 1,4 - Dichlorobenzene 2 - Methyl pentane Tetrachloroethylene 4 1,3,5 - Trimethyl benzene Hexane m / p Xylene 5 Ethylbenzene 2 - Methyl hexane/benzene Ethylbenzene 6 Styrene Heptane Trichloroethylene 7 1,1,1 Trichloroethane Toluene Toluene 8 m / p - Xylene m / p Xylene 1,1,1 Trichloroethane 9 Benzene o - Xylene Dichlorobenzenes 10 Cyclohexane Decane Styrene 11 Naphthalene Trimethyl benzene Undecane 12 Limonene Limonene Dodecane 13 Biphenyl Octane 14 1,2,3 - Trimethyl benzene 15 m / p - Ethyl toluene

a Present work. b Bernhard et al . (1995) . c Holcomb and Seabrook (1995) .

Figure 10.1 Toluene and benzene concentrations measured in Singapore and European buildings.

10.3 Apportionment of VOCs Source Strengths in Five Buildings 221

10.2.2 Health Effects Caused by VOC s in S ingapore Buildings

Many common VOCs have well - documented health effects at elevated levels. In most indoor environments however, the exposure is described as low since con-centrations are negligible in comparison to occupational limit values. Sensory responses to VOCs also include odor response, nasal irritation (pungency) and eye irritation ( Devos et al ., 1990 and Cometto - Mu iz and Cain, 1990 ). The maximum concentrations determined in Singapore buildings were compared with some of the common health and comfort guidelines. The values are presented in Table 10.5 . It is observed that the maximum concentrations for the majority of the target compounds are within recommended guidelines. Levels are very low as compared to threshold limit values related to health and irritation.

10.2.3 Possible Sources

Using principal components analysis, VOCs can be grouped into common factors that display similar source emission profi les. From the main database of eight buildings, we have grouped different VOCs into components (sources). Table 10.6 shows the extracted components using the PCA analysis of the VOC database (Zuraimi et al ., 2006 ).

Component 1 in Singapore buildings was correlated with compounds associated with humans and their activities. Human effl uents have been reported to contain isoprene (Ellin et al ., 1974 ) while tetrachloroethylene is a VOC found in dry - cleaned clothes worn by building occupants (Wallace, Pellizzari and Wendel, 1991 ) or from the use of consumer products (Sack et al ., 1992 ). Tetradecane, benzaldehyde, o - xylene, naphthalene are emissions from dry process photocopiers (Leovic et al ., 1996 ). Component 2 with high loadings of n - decane, n - undecane, toluene, styrene, n - nonane, 1,2,4 - trimethyl benzene probably refl ects the emissions of carpets and vinyl fl oorings (Yu and Crump, 1998 ). Component 3 was primarily correlated with heptane and methylcyclopentane, which could be due to the emissions of water - based paints. Finally, component 4 was associated with 2 - methylpentane, hexane, cyclohexane, methylcyclohexane and limonene, which is refl ective of the emissions of air fresheners and cleaning products (Sack et al ., 1992 ).

10.3 Apportionment of VOC s Source Strengths in Five Buildings

10.3.1 Area - Specifi c Emission Rates of VOC s

Using the concentration derived from the fi eld measurements, emission rates of VOCs in actual buildings can be calculated. This will help to determine if


Table 10.5 Maximum concentrations of the target compounds in Singapore buildings compared with health and comfort thresholds.

Compound Concentrations ( μ gm − 3 )

TLV a 0.03 × RD 50 (1/40 TLV) b Odor threshold c

Irritation threshold d

Benzene 444.1 31 566 326 099 e 810 12 020 9 000 000 Toluene 3209 375 000 380 400 f 9 380 5 890 750 000 o - xylene 255.8 434 000 190 770 f 4 769 3 800 na g m / p - xylene 499.9 434 000 h 172 320 f , h 4 308 h 1 410 i / 2 140 h na g Naphthalene 744.8 na g 67 143 d 1 679 na g na g 2 - methypentane 31.3 na g 256 544 e 6 414 na g na g 3 - methylpentane 28.8 na g 256 544 e 6 414 na g na g n - hexane 210.2 na g 256 544 e 6 414 79 430 1 800 000 n - heptane 150.6 1 618 800 1 914 030 f 40 470 40 740 na g n - decane 294.0 na g 41 420 e 1 036 4 370 na g n - undecane 278.5 na g 25 457 d 636 na g na g n - dodecane 160.2 na g 15 504 e 388 14 450 na g n - tetradecane 97.9 na g 5 644 e 141 na g na g Isoprene 60.9 na g 429 143 e 10 729 na g na g Ethanol 61.6 1 880 000 799 560 f 19 989 54 950 16 969 000 1 - butanol 18.7 147 716 115 050 f 2 876 1 510 3 334 600 2 - ethyl 1 - hexanol 244.4 na g 7 014 175 1 320 na g Methylcyclopentane 75.0 na g 272 571 e 6 814 na g na g Methylcyclohexane 181.5 1 600 000 177 788 e 4 445 na g na g 1,1,1 - trichlorethane 306.1 1 950 000 59 061 e 1 477 125 890 na g 2 - propanone 216.5 1 335 510 1 670 430 f 33 388 34 670 283 681 700 Benzaldehyde 184.9 na g 43 260 f 1 082 190 20 010 Limonene 278.9 na g 135 570 j 3 389 2 450 na g

a Threshold limit values from NIOSH (1990) . b Adjusted TLVs following the procedure of Nielsen, Hansen and Wolkoff (1997) . c Odor thresholds from Devos et al . (1990) . d Irritation thresholds from Cometto - Mu iz and Cain (1990) . e RD 50 values estimated using equation log RD 50 = 4.941 - (0.018 × molecular weight) as proposed by Alarie,

Nielsen and Schaper (2000) . f Mouse RD 50 values are from Schaper (1993) . g na = not available. h values are for p - xylene . i value is for m - xylene . j Mouse RD 50 value from Kasanen et al . (1998) .

ventilation can be used effectively to control VOC levels in buildings. Also, appor-tioning of VOC sources will allow one to concentrate on the most important source of VOCs that can targeted for control. Using a one - compartment mass balance model (Tham, Zuraimi and Sekhar, 2004 ), the area - specifi c emission rates within each building can be estimated. Assumptions for the use of the equation is that the indoor air of the buildings was perfectly mixed and under near steady - state conditions. Losses of VOCs through sink effects were ignored which translates the


Table 10.6 Principal component analysis of VOC s in eight Singapore buildings.

Rotated component matrix a

Compound Component

1 2 3 4

n - Tetradecane 0.870 Isoprene 0.731 Benzene 0.589 o - Xylene 0.735 Naphthalene 0.822 Tetrachloroethylene 0.680 Benzaldehyde 0.853 n - Nonane 0.781 n - Decane 0.904 n - Undecane 0.852 Toluene 0.823 Styrene 0.838 1,2,4 - Trimethyl benzene 0.780 n - Heptane 0.862 Methylcyclopentane 0.911 Cyclohexane 0.947 2 - Methyl pentane 0.769 n - Hexane 0.801 Methylcyclohexane 0.849 Limonene 0.725

a Extraction method: Principal component analysis; Rotation method: Varimax with Kaiser normalization. Rotation converged in 6 iterations. The analysis identifi ed 4 factors that accounted for 68% of variation between samples. After rotation factors 1 – 4 accounted for 21%, 20%, 14% and 14% of the variance respectively.

calculated emission rates calculated to net rates. The mass balanced equation to compute SER in μ gm − 2 h − 1 :

SER =−( )NV C C

Ai o (10.1)

Where C i and C o are the indoor and outdoor concentrations ( μ gm − 2 ), N is the air exchange rate (h − 1 ), A and V is the fl oor area (m − 2 ) and volume (m − 3 ) of the buildings studied. Ventilation studies conducted within buildings in Singapore were performed using sulfur hexafl uoride as a tracer gas and observing the decay in concentrations at the indoor locations using a multi - gas monitor based on IR PAS.

Similar to the indoor concentrations, the emission rates were highest among the aromatics, as shown in Table 10.7 (Zuraimi et al ., 2006 ). An important note


Table 10.7 Area specifi c emission rates of VOC s .

Compound Emission rates ( μ gm − 2 h − 1 )

Mean Standard deviation Range

Alkanes 2 - methypentane 13.6 9.9 3.8 – 31.3 3 - methylpentane 13.7 12.7 1.3 – 28.8 n - hexane 66.1 85.0 11.4 – 210.2 n - heptane 30.9 49.3 6.2 – 150.6 n - decane 70.4 102.4 2.4 – 294.0 n - undecane 69.9 94.4 7.8 – 278.5 n - dodecane 53.6 58.5 3.6 – 160.2 n - tetradecane 99.2 110.4 7.2 – 306.1

Aromatics Benzene 200.3 145.1 71.1 – 444.1 Toluene 393.0 508.6 7.7 – 3209 o - xylene 83.7 41.9 39.7 – 255.8 m / p - xylene 207.4 163.5 0.0 – 499.9 Naphthalene 310.8 217.8 104.5 – 744.8

Alcohols Ethanol 41.9 27.8 22.2 – 61.6 1 - butanol 12.8 6.6 5.7 – 18.7 2 - ethyl 1 - hexanol 57.5 102.8 1.8 – 587.2

Cyclic Alkanes Methylcyclopentane 19.2 31.4 0.3 – 75.0 Methylcyclohexane 68.9 67.6 8.3 – 181.5

Carbonyls 2 - propanone 54.2 91.7 0.1 – 216.5 Benzaldehyde 61.8 63.3 7.2 – 184.9

Others 1,1,1 - trichlorethane 45.7 114.8 1.7 – 306.1 Isoprene 22.9 16.4 9.2 – 60.9 Limonene 122.1 98.9 13.5 – 278.9

to consider was the high variability in the emission rates levels of the abundant compounds such as the xylene isomers, 1,3,5 - trimethyl benzene, toluene, 1,1,1 - trichloroethane, naphthalene, limonene and benzene across the eight buildings. More than half of these compounds have standard deviations comparable or slightly less than their mean with wide ranges. Indeed, this has an important impact from the perspectives of ventilation as a strategy to reduce indoor concen-


Figure 10.2 Emission rates of benzene and toluene in different countries.

tration levels effectively (see below). The emission rates of benzene and toluene, when compared to other countries, are also signifi cantly elevated (Figure 10.2 ).

10.3.2 Source Apportionment of VOC Sources

In the indoor environment, sources that can contribute to the VOC levels include building materials, occupants and their activities, ventilation systems and outdoor vehicular combustibles (Daisey et al ., 1994 ; Molhave and Thorsten, 1991 ). Com-putations of indoor – outdoor ratios are frequently used to conclude indoor or outdoor sources (Daisey et al ., 1994 ). Generally, derived fi ndings originated from these samples collected when the ventilation systems were in operation. It is dif-fi cult to determine the major source of VOCs for the buildings studied because contributions from building materials, occupants and their activities, ventilation systems and outdoors are all present together.

Sampling VOCs at different sampling periods and comparing the SER results within these periods can help apportion the sources of VOCs (Zuraimi, Tham and Sekhar, 2003, 2004 ). The indoor air of a subset of the eight buildings ( n = 5) was sampled in three cycles – cycle (1) during ventilation systems operation and normal occupancy, cycle (2) when the ventilation system remains in operation with no occupancy and cycle (3) after the ventilation system has been shut down and occupants has left the premises. Under normal occupancy when the ventilation


system is in operation, the VOC sources that contribute to the levels are those from building materials, occupants and their activities and ventilation system. The area - specifi c emission rates labeled as SERT are to indicate the ‘ total ’ contribution of all the sources. When the ventilation system is off, the VOC source that con-tributes to the levels is from the building materials only. Here, the infi ltration rates and the concentrations obtained after shutting the system are used to obtain the area - specifi c emission rates due to building materials, SERV. The emission rates determined when there are no occupants with ventilation system running SERBV are subtracted from SERT to give the area - specifi c emission rates due to occupants and their activities, SEROA. Subtracting SERBV from SERB would give us the area - specifi c emission rates from the ventilation systems designated as SERV.

The major source contributor of individual VOCs comes from the occupants and their activities (for TVOC, the percentage relative contribution is 37.3%). The percentage relative contribution for all VOCs ranges from 29.9% (1,1,1 - trichloro-ethane) to 79.7% (1 - butanol) with a mean of 51.1%. The next signifi cant contribu-tor comes from ventilation systems (for TVOC, the percentage relative contribution is 39.0%) where the percentage relative contribution for all VOCs ranging from 7.5% (1 - butanol) to 48.6% (1,3,5 - trimethylbenzene) with a mean of 30.1% Finally, the lowest contributor of VOCs comes from the building materials (for TVOC, the percentage relative contribution is 23.7%). The percentage relative contribution for all VOCs ranges from 3.3% (benzene) to 39.4% (1,1,1 - trichloroethane) with a mean of 17.1% There appears to be a general increasing trend for the contributor of individuals VOCs sources starting from building materials, then ventilation systems and fi nally occupants and their activities. Figure 10.3 shows the pie chart

Figure 10.3 Source apportionments of alkanes and aromatics.

10.4 Effects of Typical Ventilation Operations on TVOC Levels 227

of the contributions of different sources of the two major VOC classes – the alkanes and aromatics.

10.4 Effects of Typical Ventilation Operations on TVOC Levels

Owing to the high energy penalty associated with the treatment of outdoor air in hot and humid climates, especially in the context of dehumidifying performance for all - year air - conditioning, it is quite common to employ high recirculation rates in the design and operation of ventilation and air - conditioning systems. Studies conducted in Singapore have shown that typical outdoor air change rates in air - conditioned buildings are of the order of 0.8 – 1.2 air change per hour ( ACH ) with an average of 1 ACH in most cases (Sekhar et al ., 2003a ; Sekhar, Tham and Cheong, 2002 ). These ACH values are again observed to be associated with indoor CO 2 levels of around 800 – 1000 ppm. A threshold limit value of 1000 pm is usually recommended by most IAQ Standards and Guidelines (ASHRAE Standard 62.1, 2004 ; ENV, 1996 ). As the ventilation system design of commercial and offi ce build-ings is usually based on human occupancy and, consequently, based on maintain-ing the CO 2 levels lower than 1000 ppm, it is interesting to assess the impact of these ventilation provisions on other indoor contaminants, such as TVOCs.

Carbon dioxide and TVOC measurements obtained from studies conducted in two different buildings, A and B, located in the eastern part of Singapore are now presented and discussed. Building A is located close to the airport and is a large foot - print medium rise (about 10 storeys) offi ce - cum - industrial building, in which most of the occupied zones of the building are centrally air - conditioned. A sample of the CO 2 and TVOC plots for selected sampling locations in one of the three different zones (A1, A2 and A3) is shown in Figure 10.4 . Each of these three zones is associated with its own AHU. These three zones are spread across the 6th, 7th and 8th fl oors of the building. While A1 and A2 may be considered to represent zones which are associated purely with offi ce work, A3 also includes some pro-cesses involving chemicals. It is seen from Figure 10.4 that CO 2 levels monitored for over 3 days are about 750 ppm and with this level of ventilation, the correspond-ing indoor TVOC levels are in the range of 1.5 – 1.8 ppm during normal operating hours. The recommended indoor TVOC levels is 3 ppm (ENV, 1996 ). There is a slight build - up of TVOC levels overnight or during the system shut - down period. A2 has a similar trend with CO 2 levels of around 700 ppm and the associated TVOC levels about 2 ppm. However, A3 indicates a slightly different trend, in which for similar CO 2 and ventilation levels, the TVOC levels are in the range of 2 – 4 ppm. The higher TVOC levels during normal operation are attributed to certain pro-cesses involving chemicals. It is also to be noted that the TVOC levels during system shut down period is below 2 ppm.

Building B is a nine - storey centrally air - conditioned building involving mainly offi ce work. Training is the other kind of activity performed in this building. The CO 2 and TVOC plots for selected sampling locations in one of the three different


zones (B1, B2 and B3) are shown in Figure 10.5 . Each of these three zones is again associated with its own AHU. In Figure 10.5 the CO 2 level ranges between 600 and 700 ppm, while the TVOC level ranges between 1.5 and 2 ppm during the corresponding operating hours. In B2, it is observed that there is a spatial variation of CO 2 ranging between 600 and 950 ppm and that the TVOC level ranges between 1 and 1.2 ppm during the same period. In B3, it is seen that there is a spatial varia-tion of CO 2 ranging between 800 and 1200 ppm and that the TVOC level is around

Figure 10.4 CO 2 and TVOC plots in sampling locations in zone A1 of building A.

10.4 Effects of Typical Ventilation Operations on TVOC Levels 229

Figure 10.5 CO 2 and TVOC plots in sampling locations in zone B2 of building B.

2.5 ppm during the same period. However, some sporadic peaks of 5 and 20 ppm are also observed in one location, which is attributable to a specifi c source of chemical contamination at that location.

In all the zones in both the buildings, it is seen that the build - up of TVOC levels during system shut - down periods are quite low and the introduction of ventilation is often able to dilute the TVOC levels to a reasonably acceptable level.


10.5 Effect of Purging on Indoor TVOC Levels

Depending on the type of building materials and internal fi nishes used as well as the presence of other sources of chemical pollutants in a building, it is reasonable to expect a build - up of TVOC levels overnight during system shut - down periods. This is particularly valid in high - rise commercial, offi ce and other buildings that are designed to be air - tight and for all - year comfort air - conditioning. In such situ-ations, any attempt to ‘ pull down ’ the TVOC level at the start of the air - condition-ing system operation will be of considerable interest from an IAQ perspective. This is akin to the notion of ‘ pull down ’ of cooling loads in such buildings, where the AHU is started about 0.5 – 1.0 hour before normal occupancy in the morning. Purging is an operation to achieve a rapid reduction of overnight build - up of indoor TVOC levels and the fundamental principle is that of fl ushing the indoor pollutants with clean outdoor air. It is apparent that the quality of outdoor air is absolutely crucial to yield any sort of positive effect from IAQ considerations. It is also essential that the pollutant(s) being purged with outdoor air do not exist in the outdoor air at higher than indoor levels. In other words, the outdoor air, typi-cally, should not be a primary source of any of the indoor pollutants being purged.

A purging study was conducted in a high - rise offi ce building in Singapore in 1997, details of which are now discussed (Sekhar et al ., 2003a ).

10.5.1 Purging System

The most common type of AHU room design employed in high - rise tropical build-ings has the following features:

• A fresh air intake, either drawn directly from outside through an opening in the exterior wall of the AHU room or through a centralized fresh air duct. Fresh air provision is typically about 10% of total supply air fl ow.

• A return air opening either completely ducted from the occupied zones or through a short length of duct from the ceiling plenum that dumps all the return air into the AHU room. This would typically constitute 90% of the total supply air fl ow.

• The AHU room usually serves as a mixed air plenum.

• There is usually no provision for a ducted exhaust and whatever exhaust occurs would be through the lift core and toilet exhaust paths.

A typical AHU fi tted with a purging system is presented in Figure 10.6 . The addi-tional equipment includes an axial fan, air - duct connections, and motorized shut - off dampers and a control system which is part of the building automation system ( BAS ). Before each purging operation, the fresh air and purging dampers would be opened and the return air damper would be shut. The AHU and the purging

10.5 Effect of Purging on Indoor TVOC Levels 231

fan would be switched on. This effectively supplies fresh air into the zone while the return air is discharged from the building.

In this case study building, the locations of the fresh air damper and purged air exhaust were on different facades of the building to minimize short - circuiting of the exhaust air. As the chillers of the air - conditioning system were not operated during purging, there was no wastage of energy. Purging was carried out each morning between 6 a.m. to 6.10 a.m. The purpose was to make use of the fresher morning air to replace the contaminated air within the building. An arbitrarily determined 10 - minute purging duration was used based on the capacity of the fl ushing fan which was able to supply 6 air changes per hour. The 10 - minute purging time meant that the fan would provide 1 air change during the purging operation.

10.5.2 Building Characteristics

The building selected for the study was a 22 - fl oor offi ce building located in the central business district of Singapore. The HVAC system design for this building included a purging system in all the AHUs. The occupants in the fl oors that were selected for investigation during the study were engaged mainly in building ser-vices design and consultancy services, with considerable output in terms of draw-ings and computer printouts that are common sources of indoor pollutants. It is pertinent to note that a major construction activity adjacent to this building and hazy ambient conditions prevailed during the period of this study in 1997. The average pollutant standard index ( PSI ) readings for the month of September 1997 in which the measurements were carried out ranged between 50 and 100.

Each fl oor of the building was divided into 2 air - conditioned zones, which were served by their respective AHUs, located at each end of the fl oor. Linear diffusers were employed for supply air and the return air system was based on the ceiling plenum concept. The fl oor plans for the 4th, 11th and 17th fl oors, in which the studies were carried out, are quite similar and a schematic representation is pro-vided in Figure 10.7 .

Figure 10.6 Air handling unit integrated with a purging system.

232

10 Effect of Ventilation on V

OC

s in Indoor Air

Figure 10.7 Case study of the purging system: typical fl oor plan, air handling units and sampling points.


For continuous Monitoring of each AHU - related fl oor areas in each fl oor, results of pollutant concentration of fi ve sampling points were taken, four of the points within the air - conditioned zone marked 1, 2, 3 and 4 (set - up AHU ‘ A ’ or ‘ B ’ ) respectively. The fi fth point was the outdoor air (point 6). The fresh air intake was chosen to represent the outdoor air sample.

10.5.3 Purging Measurements

The investigations were performed with a multi - gas monitor and a multiple sampler and doser, capable of continuous real - time measurements in the building. The sampling of various sampling points is achieved by means of a physical link consisting of three millimeters diameter plastic tubes, between the various points of the building and the centrally located instruments. The tubes were suspended along the ceiling level and dropped to a height of between 1.2 to 1.5 meters above the fl oor and terminated with a 5 mm size fi lter to eliminate dust. The sampling interval was typically about 15 minutes for continuous monitoring.

The measurement of IAQ parameters on a continuous monitoring mode included CO 2 , carbon monoxide (CO), TVOC in toluene equivalents and formal-dehyde (HCHO). Figures 10.8 – 10.13 represent the measured TVOC levels at the various sampling locations distributed through the height of the building, in which the effect of the early morning purging operation is investigated.

The average indoor TVOC concentration levels ranged between 1.2 and 1.6 ppm during normal occupied hours, which is within the recommended TLV of 3 ppm (ENV, 1996 ). During system shutdown periods in the night, the indoor TVOC levels rise in excess of 3 ppm as expected. It was also observed that the outdoor TVOC levels were generally lower than those indoors and typically ranged between

Figure 10.8 Sample TVOC level in 4th fl oor location served by AHU ‘ A ’ .


Figure 10.9 Sample TVOC level in 4th fl oor location served by AHU ‘ B ’ .


1 and 2.3 ppm. However, the positive effect of early morning purging coinciding with typically lower values of outdoor TVOC levels was observed, as seen in Figures 10.8 – 10.13 , when the TVOC level drops almost instantaneously to a much lower level of around 1.2 ppm.

It is important to note that there is a multitude of pollutants that are typically classifi ed as VOCs (Wolkoff et al ., 1998 ). There is, therefore, a necessity to identify




the specifi c pollutant desired to be removed, and with suitable instrumentation, ascertain the effectiveness of the purging system in removing such pollutants. As the monitoring equipment used in this study measures the total photo ioniz-able compounds with reference to toluene, it was not possible to determine the extent to which purging can dilute the specifi c volatile organic compounds of concern.


10.6 Summary

Offi ce buildings in the tropics rely on air conditioning to provide suitable indoor climate for productive work. As energy prices become a larger cost consideration in facility management, the effective use of a comprehensive strategy to mitigate the adverse effects of pollution from building materials, humans and human - related activities need to be identifi ed and remedied. The experience in Singapore indicates that an understanding of the VOC and CO 2 profi les is necessary as a fi rst step, before apportionment of the sources to indicate where mitigating measures should be deployed. The effects of ventilation, and in particular of purging, should be relied upon to reduce the exposure risks of occupants. Such information is useful initial step for the facility manager, in consultation with the building occu-pants, to consider strategies for the improvement of indoor air quality.

Other considerations which should also be taken into consideration, to provide a total approach to indoor air quality management, would include fi ltration and air - cleaning technologies (Tham, Zuraimi and Sekhar, 2004 ), types of ventilation strategies (such as personalized ventilation, displacement ventilation and their hybrids). More recent fi ndings on the acceptability of tropically acclimatized sub-jects to slight air movement (Gong et al. , 2006 ) may also potentially be used in innovative ventilation approaches that not only provide energy effi cient ventila-tion for thermal comfort, but also provide protection by minimizing exposures to agents which may otherwise be transmitted through total air conditioning modes.


References 237

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Occurrence of Semi - Volatile Organic Compounds in the Indoor Environment Werner Butte

11

11.1 Introduction

People living in countries of moderate climate are espected to spend most of their time indoors. The National Human Activity Pattern Survey ( NHAPS ), conducted in the USA, recorded that adults spend an average of 87% of their time in enclosed buildings and about 6% in enclosed vehicles (Klepeis et al ., 2001 ). Citizens of the Federal Republic of Germany, depending on season and vocational activity, stay indoors about 85% of their time, thereof about 15 hours in their homes (Schulz et al ., 1999 ). This is in accordance with time – activity studies reported recently (Brasche and Bischof, 2005 ). Thus pollutants in the indoor environment may be signifi cant sources of exposure. A classifi cation of organic indoor contaminants, according to their volatility, was given by a WHO working group (WHO, 1989 ). This group initiated the common practice of classifying chemicals according to boiling points.

VVOCs, (b.p.: < 0 to 50 – 100 ° C) and VOCs, (b.p.: 50 – 100 to 240 – 260 ° C) are transitory and predominantly found in air. Organic compounds of lower volatility, that is, SVOCs (b.p.: 240 – 260 to 380 – 400 ° C) are present in air as well as in dust, whereas POM (b.p.: > 380 ° C) is part of the dust indoors. Analyses of SVOCs in indoor air and house dust are a measure of indoor contamination but may also provide valuable information for the assessment of human indoor exposure.

This review summarizes the recent literature on the occurrence of SVOCs in the indoor environment with the main focus on anthropogenic (man - made) com-pounds other than biocides. Aspects of exposure of the occupants will also be discussed.


239

240 11 Occurrence of Semi-Volatile Organic Compounds in the Indoor Environment

11.2 Concentrations of SVOC s in Indoor Air and House Dust

11.2.1 Phenols and Their Derivatives (Other than Biocides)

Phenolic compounds are used in commercial or consumer products or building materials (Rudel et al ., 2001 ), especially ethoxylated alkylphenols of octylphenol and nonylphenol, which are widely used in surfactants (Ying, Williams and Kookana, 2002 ). They are known as endocrine disrupting compound s ( EDC ) as they bear hormonally active properties. Other EDCs found indoors include phthal-ates (Section 11.2.7 ), certain pesticides, organotin compounds (Section 11.2.5 ) and polybrominated diphenyl ethers (Section 11.2.8 ) (Rudel et al ., 2001, 2003 ).

Rudel et al . (2001) reported concentrations of nonylphenol and its mono - and di - ethoxylates up to 14 mg kg − 1 in house dust for a 7 - sample pilot study in the USA. Concentrations of ethoxylated octylphenols did not exceed 5 mg kg − 1 . Of the more than 30 phenols analyzed only 4 - nonylphenol was found in air with concentrations up to 0.118 μ g m − 3 . Wilson, Chuang and Lyu (2001) reported the nonylphenol content of dust samples from 10 child care centers (USA) to be 4.16 – 13.8 mg kg − 1 ; the Bisphenol A content to be 1.04 – 4.51 mg kg − 1 , respectively. In their study con-centrations in air amounted to 0.052 – 0.527 μ g m − 3 for nonylphenol and up to 0.0018 μ g m − 3 for Bisphenol A.

Results for endocrine disrupting phenols in house dust as reported by Rudel et al . (2003) and Butte et al . (2001) are compiled in Table 11.1 . Rudel et al . (2003) sampled indoor air in 120 homes as well. They analyzed both house dust and air for 89 organic chemicals identifi ed as EDCs. The most abundant compounds in air included 4 - nonylphenol and 4 - t - butylphenol with typical concentrations in the range of 0.050 – 1.500 μ g m − 3 . Saito, Onuki and Seto (2004) collected air samples from houses, offi ces, and outdoor points. 4 - t - butylphenol, 4 - t - octylphenol and 4 - nonylphenol were detected in both indoor and outdoor air. Concentrations and detection frequencies were higher in indoor air than outdoor air. The maximum levels of 4 - t - butylphenol, 4 - t - octylphenol and 4 - nonylphenol in indoor air were 0.387, 0.0457 and 0.680 μ g m − 3 , respectively. 4 - t - butylphenol and 4 - nonylphenol were detected with high frequencies (more than 97%) in the indoor air samples. Wilson, Chuang and Lyu (2001) reported a mean of 0.0007 μ g m − 3 Bisphenol A in the air of 10 child care centers and a mean of 0.203 μ g m − 3 for the sum of nonyl-phenol and its ethoxylates.

Kirchner and Pernak (2004) investigated building material and analyzed indoor air for phenolic compounds in several buildings in East Berlin, Germany. Their work was done because of complaints of users, construction workers and the construction management about odor annoyance and health problems like head-ache and discomfort. Phenolic compounds in indoor air summed up to concentra-tions between 15 and 105 μ g m − 3 with 4 - ethylphenol and 2.4 - dimethylphenol as most abundant. Although presumably not all substances responsible for the malodor may have been found, the main sources were identifi ed. These were

11.2 Concentrations of SVOCs in Indoor Air and House Dust 241

Table 11.1 Concentrations of endocrine disrupting phenols in house dust (mg kg − 1 ).

Phenol Median 95th Perc. Maximum Reference

Bisphenol A 3.4 9.2 Butte et al . (2001) a Bisphenol A 0.821 17.6 Rudel et al . (2003) b t - Butylhydroxyanisole ≤ 0.1 2.0 Butte et al . (2001) a 2.4 - Dihydroxybenzophenon 0.515 9.36 Rudel et al . (2003) b techn. Nonylphenol 6.2 18 Butte et al . (2001) a 4 - Nonylphenol 2.58 8.68 Rudel et al . (2003) b Nonylphenolmonoethoxylate 3.36 15.6 Rudel et al . (2003) b Nonylphenoldiethoxylate 5.33 49.3 Rudel et al . (2003) b Nonylphenolethoxycarboxylate 2.12 9.45 Rudel et al . (2003) b n - Octylphenol ≤ 0.1 1.5 Butte et al . (2001) a Octylphenolmonoethoxylate 0.13 1.99 Rudel et al . (2003) b Octylphenoldiethoxylate 0.306 2.12 Rudel et al . (2003) b t - Octylphenol 0.3 0.86 Butte et al . (2001) a

a From homes in Northern Germany ( n = 281). b From homes on Cape Cod, USA ( n = 120, only phenols with medians greater than the

detection limit are listed).

mainly an adhesive manufactured on phenol resin basis, but also the use of phenol resin - based primers, and disinfectants added to cleaning agents.

11.2.2 Biocides

Biocides (pesticides) are chemicals to ward off pests. Biocides found indoors have mostly been used to fi ght against fungi or insects (cockroaches, fl eas, fl ies, mos-quitoes, termites etc.) and are present as pest protective agents on wood or in textiles. They were mostly applied in the homes by the occupants themselves. Data collected from 238 families in Missouri during telephone interviews showed that nearly all families (97.8%) used biocides at least once and more than two - thirds used pesticides more than fi ve times per year (Davis, Brownson and Garcia, 1992 ). The statements obtained by interviews were confi rmed by pesticide residues in the dust of the households. The situation in Germany seems to be similar to that observed in the USA. Biocides are used more often than assumed. Walker, Hostrup and Butte (1999) reported only two out of 336 house dust samples to be free from 40 target pesticides.

Insecticides and fungicides used in interiors are the same as those used in agriculture and forestry, but herbicides are seldom applied indoors (Butte, Schencke and Heinzow, 2006 ). In contrast to the outdoor environment, where modern pesticides are degraded rather quickly by microorganisms, hydrolysis and UV light, biocides applied indoors tend to be persistent. Adsorbed to a dry and dark medium such as house dust, an abiotic degradation may hardly occur and


metabolism through microorganisms cannot take place because of the lack of water (Butte and Heinzow, 2002 ). Biocides belong to SVOC or POM. Semi - volatile biocides like chlorpyrifos, lindane ( γ - HCH) and pentachlorophenol are present in air as well as in house dust (Butte, 1999 ). Since vapor pressures of ‘ non - volatile ’ biocides, for example, pyrethroids, are small, their analysis in air might not be appropriate to detect an indoor contamination. For these biocides the analysis of dust is preferred.

Biocides most often found in the indoor environment are chlorinated hydrocar-bons like chlordane, DDT, dieldrin, lindane, heptachlor and methoxychlor, pyre-throids like cyfl uthrin, cypermethrin, and permethrin, organophosphates like chlorpyrifos, diazinon, dichlorvos, isofenfos, and malathion, carbamates like ben-diocarb, carbaryl and propoxur and chlorophenols like pentachlorophenol ( PCP ), chlorocresol (4 - chloro - 3 - methylphenol) and o - phenylphenol. Residues formed in house dust may vary in different countries (Butte, 2003 ), but biocides like chlor-pyrifos, DDT, methoxychlor, permethrin, pentchlorophenol and propoxur seem to be the active compounds in biocide formulations even in different continents, as they are found equally in house dust samples form Germany and the USA (Becker et al ., 2002 ; Butte, 2003 ; Camann, Colt and Zuniga, 2002 ). Concentrations of bio-cides in house dust are mostly in the milligram per kilogram range, they seldom exceed a microgram per cubic meter in indoor air.

For PCP a signifi cant correlation of concentrations in dust with those in air was observed by Schnelle - Kreis et al . (2000) . They therefore state that the measurement of PCP in house dust is a suitable screening method for the evaluation of indoor contamination. On the other hand Stolz et al . (1996) , reporting results for perme-thrin in dust samples and air, observed no linear relationship, although air samples with higher concentrations of permethrin had a tendency to high concentrations in dust as well. The data of Roinestad, Louis and Rosen (1993) suggest that very vola-tile pesticides such as dichlorvos, o - phenylphenol and chlornaphthaline are more appropriately sampled in the air. The majority of household pesticides, however, are preferably detected by dust sampling (Roinestad, Louis and Rosen, 1993 ).

Concentrations of some biocides in the indoor environment show a tendency to decrease, others increase. According to the German environmental survey s ( GerES ), performed for nearly 20 years, concentrations of PCP in house dust are declining, a tendency already reported (Butte, 2003 ). On the other hand concentra-tions of lindane, piperonylbutoxide and permethrin seem to increase (Becker et al ., 2002 ; Seifert et al ., 2000 ).

Methods to determine biocides most commonly applied in the indoor environ-ment as well as concentrations obtained have been discussed in detail in reviews by Butte (1999) , Butte and Heinzow (2002) and Butte (2004a) .

11.2.3 Musk Compounds

Synthetic musks mainly from two groups, aromatic nitro - musks (e.g., musk xylene and musk ketone) and polycyclic musks, are produced in great quantity because


of their musk - like odor, thus replacing the original musk fragrances, extracted from exocrine gland secretions of the musk ox and the musk deer. They are widely employed for cosmetics and perfumes, many kinds of cleaning, polishing and washing agents, household products and aromatic oils (Fromme et al ., 2004a ).

The occurrence of musk fragrances (polycyclic musks in particular) in room air samples and house dust from 59 apartments and 74 kindergartens in Berlin in 2000 and 2001 was reported by Fromme et al . (2004a) . Results are reviewed in Table 11.2 . Musk compounds most often detected in the indoor air and dust were HHCB [ 1.3.4.6.7.8 - hexahydro - 4.6.6.7.8.8 - hexamethylcyclopenta - (g) 2 - benzopyrane , Galaxolide ® , Abbalide ® , Pearlide ® ], AHTN [ 7 - acetyl - 1.1.3.4.4.6 - hexa-methyl - tetraline , Tonalide ® , Fixolide ® ] and musk ketone (4 - t - butyl - 2.6 - dimethyl - 3.5 - dinitroacetophenone).

Similar results regarding house dust were reported by Butte (2004b) for nitro - musks as well as for polycyclic musks. Of the 5 nitro - musks analyzed only musk xylene (1 - t - butyl - 3.5 - dimethyl - 2.4.6 - trinitrobenzene) and musk ketone were present in nearly all of the 10 samples, but concentrations never exceeded a few milligrams per kilogram. Polycyclic musks however, especially HHCB and AHTN showed rather high residues in the 35 dust samples analyzed with concentrations for HHCB up to about 0.1 g kg − 1 .

11.2.4 Organophosphates

The phosphoric acid triesters (organophosphates, not to be mistaken for organo-phosphorous pesticides) are, like phthalates, plasticizers mixed into polymers to increase fl exibility and workability. Unlike phthalates they are remarkable fl ame retardants as well. The total European production of phosphorous - base fl ame

Table 11.2 Concentrations of synthetic musks in indoor air ( μ g m − 3 ) and house dust (mg kg − 1 ) from apartments in Berlin, Germany (Fromme et al ., 2004a ).

Median 95th Perc. Maximum

Indoor air HHCB 0.101 0.245 0.299 AHTN 0.044 0.088 0.107 AHMI 0.020 0.044 0.077

House dust HHCB 0.7 5.0 11.4 AHTN 0.9 2.3 3.1 Musk ketone 0.3 7.3 47.0

HHCB: Galaxolide ® , Abbalide ® , Pearlide ® , AHTN: Tonalide ® , Fixolide ® , AHMI: Phantolide ® .


retardants reached about 100 000 metric tons in 1997 (Ingerowski, Friedle and Thumulla, 2001 ).

Organophosphates, mainly tris(2 - chloroethyl) phosphate ( TCEP ) and tris(chloroisopropyl) phosphate ( TCPP ) are present in numerous indoor sources like adhesives, coatings, lacquers, latexes, lubricants, polymers ( polyvinyl chloride ( PVC )), polymeric resins (phenolic and phenylene - oxide - based), as well as in rigid and fl exible polyurethane foam ( PUF ) (Butte, 2004c ). From these sources organo-phosphates may readily be released (Salthammer, Fuhrmann and Uhde, 2003 ). Concentrations of TCEP and TCPP in items and material of the indoor environ-ment (wood preservation coatings, polyurethane mattresses, glass fi ber wall paper, polyurethane carpet backing, soft PUF, PUF fi llers, fl oor sealing material, and coating of acoustic ceilings) have been reported by Ingerowski, Friedle and Thumulla (2001) . They may sum up to more than a hundred grams per kilogram.

Results for organophosphates in house dust (fl oor dust and passively deposited dust from buildings in Germany and Sweden) are summarized in Table 11.3 . As a consequence of the widespread use, concentrations of organophosphates in house dust may reach some thousand milligrams per kilogram. Marklund, Ander-sson and Haglund (2003) published results for the analysis of organophosphates in settled house dust from 15 indoor environments and in wipe test samples from computer screens and covers. Tris(2 - butoxyethyl) phosphate ( TBEP ) was the most abundant organophosphate followed by TCEP and TCPP. In wipe test samples from computers, triphenyl phosphate ( TPP ) proved to be the main component.

Regarding organophosphates in indoor air, no data based on large collectives are available. Thus it is impossible to give medians and ranges of concentrations as for organophosphates in house dust. Hansen et al . (2001) analyzing TCEP, TBEP, TPP and para - tricresyl phosphate (p - TCP ) in indoor air, only found TCEP. In rooms expected to bear no sources for organophosphates, TCEP concentrations were about 20 ng m − 3 (= 0.02 μ g m − 3 ); in rooms containing material or items equipped with TCEP, concentrations between 0.33 and 3.9 μ g m − 3 were observed. Ingerowski, Friedle and Thumulla (2001) published concentrations up to 6.0 μ g m − 3 for TCEP in indoor air (median: 0.01 μ g m − 3 , n = 50). Sagunski et al . (1997) reported about 3/4 of TCEP in indoor air to be gaseous, 1/4 to be particle bound. Air samples from a Swedish offi ce building, a day care center and 3 school buildings analyzed for TBEP, tributyl phosphate ( TB P), TCEP, TCPP, TPP and tris(2 - ethylhexyl) phosphate ( TEHP ) showed concentrations between 0.0004 and about 0.030 μ g m − 3 with the highest concentration for TCEP (0.25 μ g m − 3 ) (Carlsson et al ., 1997 ). Marklund, Andersson and Haglund (2005) analyzed eleven organophos-phates in indoor air from 17 domestic and occupational environments. The total amounts of organophosphates in the air samples ranged between 0.036 and 0.95 μ g m − 3 with TCPP and TCEP being the most abundant compounds (0.0004 to 0.73 μ g m − 3 ), followed by TBP with concentrations between 0.0005 and 0.12 μ g m − 3 . Public buildings tended to have about 3 – 4 times higher levels of organophosphates than domestic buildings. A correlation was observed between the TCEP concentra-tions in the air and previously reported concentrations in dust reported by

Table 11.3 Organophosphates (plasticizers and fl ame retardants) in house dust (mg kg − 1 ).

Compound Median 95. Per - centile Range N Reference

DOPP 0.29 – < 0.03 – 5.1 15 Marklund, Andersson and Haglund (2003) a

DPEHP 0.8 370 ? – 990 29 Nagorka and Ullrich (2003) b

TBEP 5.8 58 < 0.1 – 854 199 Becker et al . (2002) c 107 – 0.7 – 1310 11 Hansen et al . (2001) d

5.0 40 ? – 120 65 Kersten and Reich (2003) e 31 – 14 – 5300 15 Marklund, Andersson and Haglund (2003) a 16.1 162 ? – 210 29 Nagorka and Ullrich (2003) b

TBP 0.4 1.5 ? – 5.7 65 Kersten and Reich (2003) e 0.35 – 0.07 – 2.2 15 Marklund, Andersson and Haglund (2003) a 2.5 34.4 ? – 49.3 29 Nagorka and Ullrich (2003) b

TCEP < 0.1 1.0 < 0.1 – 6.0 199 Becker et al . (2002) c 470 – 2.5 – 2190 12 Hansen et al . (2001) d

0.6 8.4 ? – 330 1569 Haumann and Thumulla (2002) f 0.66 9.4 < 0.1 – 121 983 Ingerowski, Friedle and Thumulla (2001) g 1.6 6.2 ? – 9.5 65 Kersten and Reich (2003) e 1.4 – 0.19 – 94 15 Marklund, Andersson and Haglund (2003) a 2.5 6.33 ? – 6.84 29 Nagorka and Ullrich (2003) b 0.9 8.4 < 0.1 – 94 59 Sagunski et al . (1997) h

TCP < 0.1 0.4 < 0.1 – 80.7 199 Becker et al . (2002) c 2.2 15 ≤ 0.1 – 36 65 Kersten and Reich (2003) e

TCPP 1.0 14 ? – 470 1337 Haumann and Thumulla (2002) f 0.57 5.9 < 0.1 – 375 436 Ingerowski, Friedle and Thumulla (2001) g 1.4 12 ? – 27 63 Kersten and Reich (2003) e 2.4 – 0.47 – 73 15 Marklund, Andersson and Haglund (2003) a

TDCPP < 0.5 2.3 ? – 120 503 Haumann and Thumulla (2002) f 1.2 6.8 ≤ 0.1 – 35 62 Kersten and Reich (2003) e 1.1 – 0.2 – 67 15 Marklund, Andersson and Haglund (2003) a 1.69 12.4 ? – 18.3 29 Nagorka and Ullrich (2003) b

TEHP < 0.1 1.6 < 0.1 – 4.6 199 Becker et al . (2002) c 0.2 0.9 ≤ 0.1 – 2.0 62 Kersten and Reich (2003) e 0.16 – 0.06 – 13 15 Marklund, Andersson and Haglund (2003) a 0.80 22.4 ? – 37.5 29 Nagorka and Ullrich (2003) b

TPP 0.3 1.8 < 0.1 – 7.2 199 Becker et al . (2002) c 1 – < 1 – 220 12 Hansen et al . (2001) d 2.9 16 ? – 56 65 Kersten and Reich (2003) e 3.1 – 0.85 – 110 15 Marklund, Andersson and Haglund (2003) a 6.51 19.5 ? – 22.9 29 Nagorka and Ullrich (2003) b

a Settled house dust, Sweden. b Passively deposited house dust from residences, Germany. c German Environmental Survey III. d Public buildings with buildings parts expected to contain organophosphates, Germany. e Residences, Hamburg, Germany. f Exactly one week old ( ‘ fi ne fraction ’ ), Germany. g Exactly one week old (total dust), Germany. h Total fl oor dust, Germany. DOPP : di - n - octylphenyl phosphate , DPEHP : diphenyl 2 - ethylhexyl phosphate , TDCPP : tris(dichloroisopropyl) phosphate .


Marklund, Andersson and Haglund (2003) , but no such association was seen for the heavier and less volatile TBEP.

They occurrence of organophosphates in the indoor environment has recently been reviewed in detail by Butte (2004c) and Wensing, Uhde and Salthammer (2005) .

11.2.5 Organotin Compounds

Organotins are a large class of compounds that have been used for a variety of purposes since their discovery and initial use in the 1920s (Fromme et al ., 2005 ): They have the general formula R n SnX 3 - n , where R is an alky or aryl group, Sn the central tin atom and X a singly charged anion or anionic group. The alkyl moiety mostly is a methyl, butyl, octyl or phenyl group. Organotins have been used as heat and light stabilizers in the production of PVC and as catalysts for polyure-thane and silicone elastomers. Thus, they are present in many consumer products. But organotins have also been applied as a biocide for wood, paper, textiles, paints and some electric equipment. These possible applications may be another source for an indoor contamination (tributyl tin compounds used as antifoulings for ships and boats will rarely be present indoors).

Table 11.4 summarizes concentrations for MBT ( monobutyltin ), DBP ( dibutyl-tin ), TBT ( tributyltin ), MOT ( monooctyltin ), DOP ( dioctyltin ) and TPT ( triphenyl-tin ) as reported from the Greenpeace research laboratories (Greenpeace Nederland, 2001 ; Santillo, Johnston and Bridgen, 2001 ; Santillo et al ., 2003 ) as well as from Haumann and Thumulla (2002) , Kersten and Reich (2003) and Fromme et al . (2005) . MBT is normally the most abundant organotin in house dust (Kersten and Reich, 2003 ; Santillo, Johnston and Bridgen, 2001 ; Santillo et al ., 2003 ), but in the studies of Greenpeace Nederland (2001) , Haumann and Thumulla (2002) and Fromme et al . (2005) DBT revealed the highest abundance.

Total organic tin compounds in house dust mostly amount to some milligrams per kilogram, but they might form residues of some hundred mg kg − 1 as well (Greenpeace Nederland 2001 ). Concentrations of organotins in indoor air have not yet been published.

11.2.6 Perfl uorinated Compounds

Perfl uoroalkyl compounds have been manufactured since the 1950s.The total production of fl uorinated surfactants (anionic, cationic and neutral) was 200 t in 1979, whereas in 2000 the total production of PFOS ( perfl uorooctane sulfonate ) alone was nearly 3000 t (Shoeib et al ., 2004 ). Together with PFOA ( perfl uoroocta-noic acid ), PFOS is used in refrigerants, surfactants, fi re retardants, stain - resistant coatings for fabrics, carpets and paper and insecticides. Surface treatments, such as protection of clothing and carpets constitute the largest volume of PFOS produc-tion (Moriwaki, Takata and Arakawa, 2003 ). PFOA as well is present in several


Table 11.4 Organotin compounds in house dust (mg kg − 1 ).

Compound Median Range Reference

MBT 0.41 0.012 – 81.8 Greenpeace Nederland (2001) a 1.05 0.18 – 2.39 Santillo, Johnston and Bridgen (2001) b 0.5 0.1 – 24 Haumann and Thumulla (2002) c 1.35 0.81 – 2.80 Santillo et al . (2003) d 1.4 0.1 – 18.0 Kersten and Reich (2003) e 0.05 0.005 – 1.50 Fromme et al . (2005) f

DBT 0.25 0.005 – 130 Greenpeace Nederland (2001) a 0.464 0.172 – 0.891 Santillo, Johnston and Bridgen (2001) b 3.4 0.1 – 2200 Haumann and Thumulla (2002) c 0.52 0.16 – 1.3 Santillo et al . (2003) d 0.2 0.01 – 5.6 Kersten and Reich (2003) e 0.03 < 0.005 – 5.6 Fromme et al . (2005) f

TBT 0.009 0.005 – 2.03 Greenpeace Nederland (2001) a 0.022 0.004 – 0.047 Santillo, Johnston and Bridgen (2001) b 0.5 0.1 – 15 Haumann and Thumulla (2002) c 0.05 0.02 – 0.76 Santillo et al . (2003) d 0.03 < 0.001 – 0.2 Kersten and Reich (2003) e 0.008 < 0.005 – 0.08 Fromme et al . (2005) f

MOP 0.02 0.005 – 0.98 Greenpeace Nederland (2001) a 0.261 0.015 – 0.832 Santillo, Johnston and Bridgen (2001) b 0.35 0.08 – 1.3 Santillo et al . (2003) d 0.2 0.01 – 2.8 Kersten and Reich (2003) e 0.008 < 0.005 – 0.04 Fromme et al . (2005) f

DOT 0.02 0.005 – 2.12 Greenpeace Nederland (2001) a 0.057 0.004 – 0.140 Santillo, Johnston and Bridgen (2001) b 0.06 0.02 – 0.55 Santillo et al . (2003) d 0.01 < 0.001 – 0.2 Kersten and Reich (2003) e 0.005 < 0.005 – 0.36 Fromme et al . (2005) f

TPT < 0.001 < 0.001 – 0.07 Santillo et al . (2003) d 0.004 < 0.001 – 0.02 Kersten and Reich (2003) e

< 0.005 – Fromme et al . (2005) f

a Apartments and large buildings, Netherlands ( n = 134). b Collected in parliament buildings in 8 European countries ( n = 15). c Samples exactly 7 days old, Germany ( n = 33). d From all UK regions (partly pooled) ( n = 100). e Apartments in Hamburg, Germany ( n = 50). f Residences in Berlin, Germany ( n = 28).


consumer articles (carpeting, home textiles, latex paint, home and offi ce cleaners) (Washburn et al ., 2005 ).

PFOA, PFOS and PFHS ( perfl uorohexanoic acid ) concentrations have been reported for house dust from Japan and Canada. The concentrations of PFOS and PFOA in vacuum cleaner dust collected in Japanese homes ( n = 16) were: PFOS: median 24, range 11 – 2500 μ g kg − 1 and PFOA: median 165, range: 69 – 3700 μ g kg − 1 (Moriwaki, Takata and Arakawa, 2003 ). Similar results were found in house dust from Canadian homes ( n = 69), PFOS: median 37.8, range 2.28 – 5065 μ g kg − 1 , PFOA: median 19.7, range 1.15 – 1234 μ g kg − 1 and PFHS: median 23.1, range 2.28 – 4305 μ g kg − 1 (Kubwabo et al ., 2005 ). The data revealed a correlation between the concentrations of perfl uorinated compounds and the percentage of carpeting in the house; older houses tended to have less carpeting, hence levels of perfl uori-nated compounds in their dust were lower (Kubwabo et al ., 2005 ).

Concentrations of the perfl uorinated sulfonamides MeFOSE ( N - methylperfl uo-rooctane sulfonamidoethanol ) and EtFOSE ( N - ethylperfl uorooctane sulfonamido-ethanol ) in indoor air and dust were reported by Shoeib et al . (2005) . These sulfonamides are precursors of PFOS. Residues in house dust ( n = 66) were: MeFOSE: geometric mean: 113, range 3.3 – 8860 μ g kg − 1 and EtFOSE: geometric mean: 138, range 1.4 – 75 440 μ g kg − 1 ; concentrations in indoor air were: geometric mean: 1.49, range 0.366 – 8.19 ng m − 3 and EtFOSE: geometric mean: 0.744, range 0.227 – 7.740 ng m − 3 . Although there was a considerable scatter in the data, a good agreement with higher dust to higher indoor air concentrations was observed (Shoeib et al ., 2005 ).

11.2.7 Phthalates

Dialkyl and alkyl arylesters of phthalic acid, that is, phthalates, are industrial chemicals with a wide range of applications. Phthalates are primarily used as plasticizers mixed into polymers like PVC to increase fl exibility and workability. Di(2 - ethylhexyl) phthalate ( DEHP ), di(isononyl) phthalate ( DINP ) and di(isodecyl) phthalate ( DIDP ) are the general purpose plasticizers for PVC. Di( n - butyl) phthal-ate ( DBP ) and butylbenyzl phthalate ( BBzP ) are fast - fusing plasticizers for PVC as well and are mostly used in combination with DEHP. BBzP is further present in cosmetics like hair sprays. DBP and di - i - butyl phthalate ( DIBP ) are also appropri-ate plasticizers for nitrocellulose lacquers (nail polish). Phthalates are further reported to be present in textiles like cotton diapers, bathrobes, T - shirts, upholstery fabric, and carpeted fl oor. Modern electronic goods such as TV sets, computers, copying machines etc. are indoor sources of phthalates as well (Butte, 2004c ). From all the indoor items phthalates may evaporate. As a consequence they are present in indoor air and have the tendency to accumulate in house dust. Dust may further contain some plasticizers form plastic material that has been abraded.

There are numerous reports on the occurrence of phthalates indoors. Results from different countries are compiled in Table 11.5 for house dust and in Table 11.6 for indoor air. Although the house dust samples were of different origin and


Table 11.5 Concentrations of phthalates in house dust (mg kg − 1 ).

Phthalate Median 95. Perc. Range N Region of dust sampling

BBzP 14.7 207 ? – 745 199 German Environmental Survey III a 135 599 < 40 – 45 549 346 Children ’ s bedrooms (Sweden) b 30.5 320 0.3 – 1 400 286 Homes in Northern Germany c 29.7 218.5 ? – 815.7 30 Apartments in Berlin, Germany ( n = 59) d 19 230 ? – 700 65 Apartments in Hamburg (Germany) e 13 180 < 2 – 460 278 German Environmental Survey on children,

2001/2002 (pilot study) f 110 (mean) – < 10 – 480 38 Dwellings in Oslo (Norway) g 24 270 < 0.7 – 510 272 Homes from all over Germany h 45.4 – 3.87 – 1 310 119 Homes on Cape Cod, USA i 67.7 (mean) – 15.1 – 175 10 Child care centers in North Carolina j

BMEP 2 8 ≤ 1 – 17 65 Homes in Hamburg (Germany) e

DBP 41.5 160 ? – 502 199 German Environmental Survey III a 150 568 < 40 – 5 446 346 Children ’ s bedrooms (Sweden) b 49 240 3.5 – 500 286 Homes in Northern Germany c 47 129.6 ? – 141.4 30 Apartments in Berlin, Germany ( n = 59) d 47 180 ? – 600 65 Apartments in Hamburg (Germany) e 29 180 < 7.6 – 740 278 German Environmental Survey on children,

2001/2002 (pilot study) f 100 (mean) 10 – 1 170 38 Dwellings in Oslo (Norway) g 87 370 < 0.7 – 1 200 272 Homes from all over Germany h 20.1 – < 24 – 352 119 Homes on Cape Cod, USA i 18.4 (mean) ? 1.58 – 46.3 10 Child care centers in North Carolina j

DIBP 22.4 130 ? – 192 199 German Environmental Survey III a 45 311 < 40 – 3 810 346 Children ’ s bedrooms (Sweden) b 34 130 1.1 – 330 286 Homes in Northern Germany c 33 78 ? – 470 65 Apartments in Hamburg (Germany) e 33 120 < 4 – 310 278 German Environmental Survey on children,

2001/2002 (pilot study) f 10 (mean) < 10 – 450 38 Dwellings in Oslo (Norway) g 1.91 – < 1 – 39.1 119 Homes on Cape Cod, USA i

DCHP ≤ 1 5 ≤ 1 – 80 62 Apartments in Hamburg (Germany) e 1.88 – < 0.8 – 62.7 101 Homes on Cape Cod, USA i

DIDP 31 340 ≤ 1 – 4 200 62 Apartments in Hamburg (Germany) e 60 170 < 2.3 – 2 600 278 German Environmental Survey on children,

2001/2002 (pilot study) f

DEHP 416 1190 ? – 7 530 199 German Environmental Survey III a 770 4069 < 40 – 40 459 346 Children ’ s bedrooms (Sweden) b 735 2600 62 – 12 000 286 Homes in Northern Germany c 3214 (mean) 7063 – 15 Danish schools k 703.4 1542 ? – 1 763 30 Apartments in Berlin, Germany ( n = 59) d 600 1600 ? – 2 700 65 Apartments in Hamburg (Germany) e 480 1700 < 41 – 5 100 278 German Environmental Survey on children,

2001/2002 (pilot study) f 640 (mean) – 110 – 2 100 38 Dwellings in Oslo (Norway) g 450 2000 < 0.7 – 8 600 272 Homes from all over Germany h 340 – 16.7 – 7700 101 Homes on Cape Cod, USA i


Phthalate Median 95. Perc. Range N Region of dust sampling

DEP 3.3 89.7 < 0.1 – 1233 199 German Environmental Survey III a < 40 115 < 40 – 2425 346 Children ’ s bedrooms (Sweden) b 6.1 159.6 ? – 632.2 30 Apartments in Berlin, Germany ( n = 59) d 5 350 ≤ 1 – 570 65 Apartments in Hamburg (Germany) e 4.8 210 2.6 – 1200 278 German Environmental Survey on children,

2001/2002 (pilot study) f 10 (mean) < 10 – 170 38 Dwellings in Oslo (Norway) g 3.1 96 < 0.5 – 310 272 Homes from all over Germany h 4.98 – < 4 – 111 119 Homes on Cape Cod, USA i

DHP 1.1 – < 0.1 – 30.6 119 Homes on Cape Cod, USA i

DIHP 18 170 < 2.9 – 530 278 German Environmental Survey on children, 2001/2002 (pilot study) f

DMP 0.2 3.7 < 0.1 – 75.8 199 German Environmental Survey III a 1.5 46.4 < 0.5 – 157.9 30 Apartments in Berlin, Germany ( n = 59) d ≤ 1 20 ≤ 1 – 64 65 Apartments in Hamburg (Germany) e 0.57 9.8 < 0.53 – 1 300 278 German Environmental Survey on children,

2001/2002 (pilot study) f 0.6 4.1 < 0.5 – 35 272 Homes from all over Germany h

DMPP 37.5 144.4 ? – 161.3 30 Apartments in Berlin, Germany ( n = 59) d

DINP 41 1930 < 40 – 40 667 346 Children ’ s bedrooms (Sweden) b 72 540 ≤ 1 – 1 000 61 Apartments in Hamburg (Germany) e 80 280 < 2.1 – 1400 278 German Environmental Survey on children,

2001/2002 (pilot study) f

DOP 4 73 ≤ 1 – 160 65 Apartments in Hamburg (Germany) e

a Becker et al . (2002) . b Bornehag et al . (2005) . c Butte et al . (2001) . d Fromme et al . (2004a) . e Kersten and Reich (2003) . f Nagorka, Scheller and Ullrich (2005) . g Ø ie et al . (1997) . h P ö hner et al . (1998) . i Rudel et al . (2003) . j Wilson, Chuang and Lyu (2001) . k Clausen et al . (2003) . BMEP : bis(2 - methoxyethyl) phthalate , DCHP : dicyclohexyl phthalate , DEP : diethyl phthalate , DHP : di( n - hexyl) phthalate , DIHP : di(isoheptyl)phthalate , DMP : dimethyl phthalate , DMPP : dimethylpropyl phthalate , DOP : di( n - octyl) phthalate .

Table 11.5 Continued


Table 11.6 Concentrations of phthalates in indoor air (ng m − 3 ).

Phthalate Median Range Location

BzBP 40 10 – 630 Personal air sampling, New York ( n = 30) a 20 < 10 – 190 Personal air sampling, Krakow ( n = 30) a 18 < 10 – 575 Apartments in Berlin, Germany ( n = 59) b

< 10 < 10 – 391 Kindergartens in Berlin, Germany ( n = 74) b – < 1.2 – 100 Pilot study in 6 contemporary Japanese houses c < 31 < 31 – 480 Homes on Cape Cod, USA ( n = 120) d 100 (mean) 11.6 – 581 Child care centers in North Carolina ( n = 10) e

DBP 400 110 – 4100 Personal air sampling, New York ( n = 30) a 2300 750 – 15 000 Personal air sampling, Krakow ( n = 30) a 829 572 – 1346 Offi ce, classroom and room in a day - care center, Denmark,

( n = 12) f 1083 ? – 5 586 Apartments in Berlin, Germany ( n = 59) b 1188 ? – 13 305 Kindergartens in Berlin, Germany ( n = 74) b – 110 – 600 Pilot study in 6 contemporary Japanese houses c 220 52 – 1100 Homes on Cape Cod, USA ( n = 120) d 420 1300 (90. percentile) Californian homes, ( n = 125) g 239 (mean) 108 – 404 Child care centers in North Carolina (USA), ( n = 10) e

DIBP 61 11 – 990 Homes on Cape Cod, USA ( n = 120) d

DCHP – < 1.2 – 170 Pilot study in 6 contemporary Japanese houses c < 2 < 2 – 280 Homes on Cape Cod, USA ( n = 120) d

DEHP 220 50 – 410 Personal air sampling, New York ( n = 30) a 370 80 – 1 100 Personal air sampling, Krakow ( n = 30) a 258 111 – 1 053 Offi ce, classroom and room in a day - care center, Denmark,

( n = 12) f 156 ? – 615 apartments in Berlin, Germany ( n = 59) b 458 ? – 2 253 Kindergartens in Berlin, Germany ( n = 74) b

– 40 – 230 Pilot study in 6 contemporary Japanese houses c 77 < 59 – 1 000 Homes on Cape Cod, USA ( n = 120) d

110 240 (90. percentile) Californian homes, ( n = 125) g

DEP 2700 1 500 – 7 100 Personal air sampling, New York ( n = 30) a 840 260 – 2 900 Personal air sampling, Krakow ( n = 30) a 643 ? – 5 481 Apartments in Berlin, Germany ( n = 59) b 353 ? – 1 263 Kindergartens in Berlin, Germany ( n = 74) b ,

– 50 – 190 Pilot study in 6 contemporary Japanese houses c 590 130 – 4 300 Homes on Cape Cod, USA ( n = 120) d

DMP 436 ? – 13 907 Apartments in Berlin, Germany ( n = 59) b 331 ? – 13 233 Kindergartens in Berlin, Germany ( n = 74) b

DMPP 459 ? – 5 887 Apartments in Berlin, Germany ( n = 59) b 505 ? – 2 659 Kindergartens in Berlin, Germany ( n = 74) b

a Adibi et al . (2003) . b Fromme et al . (2004a) . c Otake, Yoshinaga and Yanagisawa (2001) . d Rudel et al . (2003) . e Wilson, Chuang and Lyu (2001) . f Clausen, Wolkoff and Svensmark (1999) . g Sheldon et al . (1993) .


Figure 11.2 Phthalates in dust (Kersten and Reich 2003 ).

Figure 11.1 Phthalates in indoor air (Fromme et al ., 2004a ).

although different material was analyzed (total dust, dust fraction ≤ 2 mm, etc.), results are remarkably consistent. But regarding the phthalates present indoors concentrations may vary several orders of magnitude. Phthalates with a high vola-tility dominate in indoor air (Figure 11.1 ), whereas phthalates in dust represent the production volume (Butte, 2004c ), thus DEHP is the phthalate showing the highest residues in house dust (Figure 11.2 ).

Emission of phthalates form plastic material, especially PVC, and its dispersion in air as well as its uptake in dust were reported by Afshari et al . (2004) , Clausen et al . (2004) , Fujii et al . (2003) and Uhde et al . (2001) . Bornehag et al . (2005) exam-


ined associations between the concentrations of phthalates in dust and various characteristics of the home, focusing on BBzP and DEHP. For both phthalates associations between their dust concentrations and the amount of PVC used as fl ooring and wall material in the home was found. Furthermore high concentra-tions of DEHP were associated with buildings constructed before 1960. But both, BBzP and DEHP, were also found in buildings where neither PVC fl ooring nor wall covering was present.

For further information on plasticizers (organophosphates and phthalates) in the indoor environment see Butte (2004c) and Wensing, Uhde and Salthammer (2005) .

11.2.8 Polybrominated Diphenyl Ethers

Polybrominated diphenyl ether s ( PBDE s) and hexabromocyclododecane ( HBCD ) are widely used as additive fl ame retardants in a wide array of industrial and con-sumer products including electrical and electronic appliances, vehicles, lighting and wring, textiles (including carpets and other furnishing) and packing and insulating materials (especially polystyrene) (Santillo et al ., 2003 ). Thus they are present in the indoor environment and may be released indoors via volatilization or as dusts. The penta - and octabrominated mixes are now banned in most parts of Europe, and phasing out of their use has recently begun in North America (Wilford et al ., 2005 ).

The occurrence of PBDEs and HBCD in house dust from European countries and the USA is complied in Table 11.7 . The penta - and hexa - and heptabrominated congeners were frequently detected in the house dust samples, with BDE 209 (BDE = brominated diphenyl ether) as the predominant compound reaching con-centrations up to a hundred milligram per kilogram. Correlations between penta-bromo mix congener levels in dust and in air from the same homes were reported, but not for congeners of the more highly brominated mixes (Wilford et al ., 2005 ).

11.2.9 Polychlorinated Biphenyls

Polychlorinated biphenyl s ( PCB s), although banned in many countries nowadays, belong to the most ubiquitous environmental pollutants because of their longevity and persistence. An indoor contamination may result from their use as insulating materials in electrical capacitors and transformers, but they have also been added to paints, coatings and adhesives (as fl ame retardants), and they are plasticizers in all kinds of plastics and cable sheathings (Sagunski, Rosskamp and Heinrich - Hirsch, 1998 ). In Germany, a high PCB contamination indoors mainly results from their use as plasticizer in elastic rubber sealing ( ‘ Thiokol ’ ) (Benthe et al ., 1992 ).

PCBs consist of 209 congeners. Mostly the six indicator congeners 28, 52, 101, 138, 153, 180 are chosen for analysis. These indicator PCB congeners show differing


Table 11.7 Concentrations of selected polybrominated diphenyl ethers ( PBDE ) and hexabromocyclododecane ( HBCD ) in house dust (mg kg − 1 ).

PBDE Median Range Reference

BDE 28 (tri - ) 0.000 69 0.000 12 – 0.0028 Santillo, Johnston and Bridgen (2001) a 0.000 35 < 0.000 01 – 0.033 Santillo et al . (2003) b 0.003 ? – 0.55 Wilford et al . (2005) c

BDE 47 (tetra - ) 0.059 0.01 – 0.18 Santillo, Johnston and Bridgen (2001) a 0.0145 0.0033 – 1.91 Knoth et al . (2002) d – ? – 9.86 Rudel et al . (2003) e 0.0248 0.01 – 1.98 Santillo et al . (2003) b 0.30 0.021 – 33.0 Wilford et al . (2005) c

BDE 99 (penta - ) 0.050 0.01 – 0.17 Santillo, Johnston and Bridgen (2001) a 0.0159 0.0026 – 2.85 Knoth et al . (2002) d 0.304 ? – 22.5 Rudel et al . (2003) e 0.044 0.018 – 2.1 Santillo et al . (2003) b 0.43 0.019 – 60.0 Wilford et al . (2005) c

BDE 100 (penta - ) 0.012 0.0025 – 0.036 Santillo, Johnston and Bridgen (2001) a 0.0029 0.0005 – 0.162 Knoth et al . (2002) d – ? – 3.4 Rudel et al . (2003) e 0.073 0.0041 – 21.0 Wilford et al . (2005) c

BDE 153 (hexa - ) 0.015 0.0061 – 0.059 Santillo, Johnston and Bridgen (2001) a 0.0044 0.0004 – 0.42 Knoth et al . (2002) d 0.023 < 0.0001 – 0.17 Santillo et al . (2003) b 0.049 0.0031 – 25.0 Wilford et al . (2005) c

BDE 183 (hepta - ) 0.0056 0.0009 – 0.464 Knoth et al . (2002) d 0.0095 < 0.0001 – 0.087 Santillo et al . (2003) b 0.019 ? – 0.65 Wilford et al . (2005) c

BDE 209 (deca - ) 0.61 0.26 – 6.9 Santillo, Johnston and Bridgen (2001) a 7.1 3.8 – 19.9 Santillo et al . (2003) b 0.63 0.074 – 10.0 Wilford et al . (2005) c

HBCD 0.02 < 0.0025 – 1.4 Santillo, Johnston and Bridgen (2001) a 3.25 0.94 – 6.9 Santillo et al . (2003) b

a Collected in parliament buildings and an internet provider in 8 European countries ( n = 10). b From all UK regions ( n = 10). c From Ottawa, Canada ( n = 64). d From households in the Rhein - Main area, Germany ( n = 25). e From homes on Cape Cod, USA ( n = 119).


toxicity and volatility. PCB28, 52 and 101 are rather volatile; they are more likely to be found in the air. In house dust however, mainly the semi - volatile PCB138, 152 and 180 are present. Analytical results for PCB are often given as the sum of the six indicator PCBs (multiplied by a factor of 5), thus representing the ‘ total PCB ’ con-centration. Medians and 95th percentiles for the 6 PCB congeners in house dust obtained from representative collectives in Germany and the USA are listed in Table 11.8 . Although results were obtained from different fractions of house dust (2 mm in contrast to 0.15 mm) and although the dust samples origin from two different continents, the 95th percentiles are astonishingly similar (Becker et al ., 2002 ; Camann, Colt and Zuniga, 2002 ). Becker et al . (2002) reported a 95th percentile of 0.8 mg kg − 1 for the total PCB content in house dust in Germany. Similar results were obtained by P ö hner et al . (1998) , who found a 95th percentile of 1.6 mg kg − 1 ( n = 272) for the sum of nine indicator congeners (PCB28, 52, 77, 101, 126, 138, 153, 169 and 180), but more than 80% of their PCB results were below the detection limit of 0.2 mg kg − 1 .

Concentrations of PCB52, PCB 105 and PCB 153 form homes on Cape Cod (USA) were given by Rudel et al . (2003) . Median values were below the detection limit, the maximum levels were 15.7 (PCB52), 16.5 (PCB105) and 35.3 (PCB153) mg kg − 1 .

Other work on PCBs in house dust relies on 20 PCB congeners like in the studies of Wilson, Chuang and Lyu (2001, 2003) for dust and air from 5 child care centers and from homes of 9 children. Most prominent PCB congeners in dust were PCB101 and PCB110 (both pentachlorinated PCB congeners) in both studies in contrast to air where PCB10, a dichlorinated PCB congener and PCB 28 (a trichlorinated PCB congener), were most abundant.

Table 11.8 Polychlorinated biphenyls ( PCB s ) in house dust from collectives representative for Germany and the USA (mg kg − 1 ).

PCB congener Germany a USA b

Median 95th Percentile Median 95th Percentile

PCB 28 ≤ 0.02 ≤ 0.02 PCB 52 ≤ 0.02 0.03 PCB 101 ≤ 0.01 0.11 PCB138 0.01 0.24 ≤ 0.021 0.22 PCB 153 ≤ 0.02 0.24 ≤ 0.021 0.24 PCB 180 ≤ 0.02 0.17 ≤ 0.021 0.112

a Representative for Germany sieved to ≤ 2 mm ( n = 741) (Becker et al ., 2002 ). b Collected in the Detroit area, Michigan; the entire state of Iowa, Los Angeles County, and the

Seattle area, Washington sieved to ≤ 150 μ m ( n = 616) (Camann, Colt and Zuniga, 2002 ).


65 target PCB congeners in house dust were analyzed by Vorhees, Cullen and Altshul (1999) from 34 homes surrounding New Bedford Harbor, a place were highly contaminated harbor sediments had been dredged. PCB concentrations were 0.26 – 23.0 mg kg − 1 in house dust, they were correlated with PCB concentra-tions in indoor air. But house dust and yard soil concentrations were not corre-lated, suggesting that track - in may not be the only source of PCBs in house dust.

Concentrations of 12 dioxin - like PCBs, that is, non - ortho and mono - ortho PCB congeners, in 10 house dust samples from Kumagaya and Sendai city (Japan) were reported by Saito et al . (2003) . They may sum up to 30.5 ng TEQ kg − 1 ( TEQ = toxic-ity equivalent ).

Indoor air levels of coplanar polychlorinated biphenyls (mono - and non - ortho substituted PCB) for 29 sampling sites were determined by Kohler, Zennegg and Waeber (2002) . As joint sealings containing PCB were present in some rooms, total PCB concentrations in indoor air amounted to 4200 ng m − 3 . In a PCB con-taminated industrial building, total indoor air PCB levels of up to 13 000 ng m − 3 were measured. Typical PCB congeners in indoor air included PCB 28, PCB 52 and PCB 101. Concentrations of coplanar (dioxin - like) PCB were determined for six sites. The most abundant coplanar PCB congener in indoor air was PCB 118, followed by PCB 105, PCB 123, and PCB 77 in various order. Levels of coplanar PCB, expressed as TEQ, correlated well with the total indoor air PCB concentra-tion. Kohler, Zennegg and Waeber (2002) stated that a total PCB level of 1000 ng m − 3 corresponds to a concentration of coplanar PCBs of 1.2 pg TEQ m − 3 .

Liebl et al . (2004) measured the six indicator congeners (PCB 28, 52, 101, 138, 153, 180) in indoor air of a contaminated school. The sum of the six indicator congeners (times 5) ranged from 690 and 20 800 ng m − 3 (median 2044 ng m 3 ). The lower chlorinated congeners PCB 28, 52, 101 were the prevailing contaminants (medians 33, 293 and 66 ng m − 3 ).

11.2.10 Polychlorinated Dioxins and Furans

Wittsiepe et al . (1997) published results for the levels of PCDD/PCDF in house dust of 22 residential homes located in different areas (rural, urban and industrial) of Germany ( ≤ 2 mm fraction). The average level of PCDD/PCDF in ‘ normal ’ houses was 101 ng I - TE kg − 1 with a range from 7.83 – 332 ng I - TE kg − 1 ( I - TE : Inter-national tetrachlorodibenzo - dioxine equivalent ). The predominant congener was OCDD ( octachlorodibenzo - dioxin ). The average level of PCDD/PCDF in house dust collected from a contaminated area (former metal reclamation plant) was 265 ng I - TE kg − 1 . (range: 29.9 – 1050 ng I - TE kg − 1 ). This increase refers mainly to the levels of low chlorinated PCDD/PCDF. House dust collected from rooms where wood protective agents containing PCP were used showed a characteristic PCDD/PCDF congener pattern with an especially high concentration of OCDD (3000 μ g kg − 1 ).

A method for the extraction and cleanup of chlorinated dioxins and furans from house dust was published by Saito et al . (2003) using accelerated solvent extraction


(ASE), a multi - layer silica - gel cartridge for clean - up and 13 C - labeled internal stan-dards. 17 dioxins and furans as well as 12 PCB congeners, known to have dioxin - like properties, were included in their study. It was found that the average values of the dioxins in house dust from Kumagaya and Sendai ( n = 5 each), cities in Japan, were 14.0 ng TE kg − 1 (range: 7.7 – 26.0 ng TE kg − 1 , not considering the dioxin - like PCBs) in Kumagaya and 13.1 ng TQ kg − 1 (4.3 – 25.8 ng TE kg − 1 ) in Sendai city. As in the work of Wittsiepe et al . (1997) OCDD was the most prominent isomer with concentrations up to 1.5 μ g kg − 1 . The highest concentration of 2.3.7.8 - TCDD was 2 ng kg − 1 .

Hansen and Volland (2002) measured polychlorinated dioxins and furans (2.3.7.8 - TCDD, 1.2.3.7.8 - PCDD, three HxCDDs, one HpCDD and OCDD) in house dust and air of different fl oors of a building including the attics; the build-ing was expected to be contaminated by dioxins and furans. Compared to the fi ndings of Wittsiepe et al . (1997) ; Saito et al . (2003) concentrations for the dust samples from the living rooms were much higher, that is, between 0.85 and 4.02 μ g I - TE kg − 1 . The dust from the attics was even more contaminated, i.e., 63 μ g I - TE kg − 1 . Concentrations of furans exceeded those of dioxins; the octachlorinated compounds showing concentrations in the high microgram per kilogram range (OCDD: 67 μ g kg − 1 ; OCDF 137 μ g kg − 1 were the most prominent compounds). 2.3.7.8 - TCDD was only present in one of the living rooms (4 ng kg − 1 ) and in the attics (490 ng kg − 1 ). Concentrations of TCDD and TCDF in indoor air did not exceed the detection limit (0.01 pg m − 3 ), concentrations of all the other dioxins and furans analyzed ranged from 0.02 to 0.29 pg m − 3 . OCDD and OCDF showed, as in dust, the highest concentrations. Discussing their results, the authors state that there is no simple relation between dioxins and furans in air and in dust.

11.2.11 Polycyclic Aromatic Hydrocarbons

Polycyclic aromatic hydrocarbon ( PAH ) is the collective term for aromatic hydro-carbons with condensed ring systems. PAHs are formed when burning organic compounds, so they originate from combustion processes, motor vehicles, fumiga-tion, smoking, etc. With respect to the indoor environment, tar and bitumen based adhesives used for parquet fl oors are a dominant source for a contamination, as these may contain high concentrations of PAHs (Dieckow, Ullrich and Seifert, 1999 ). Benzo[a]pyrene ( BaP ), a PAH bearing carcinogenic properties, is used as the indicator PAH in Germany, since there is a signifi cant correlation between PAHs and BaP in house dust (Dieckow, Ullrich and Seifert, 1999 ). In the USA, concentrations of probable human carcinogenic PAHs (B2 PAH) are often sum-marized. Results for PAHs in house dust and indoor air from studies performed in Germany and the USA are compiled in Tables 11.9 and 11.10 .

Lewis et al . (1999) prepared a gross house dust sample by combing dust from four vacuum cleaner bags obtained from 25 middle - class homes. The composite dust was separated into seven size fractions ranging from < 4 to 500 micrometer in diameter. Ten PAHs were analyzed. All of the ten target PAHs were detected


Table 11.9 Concentrations of some polycyclic aromatic hydrocarbons ( PAH s ) in house dust from Germany and the USA (mg kg − 1 ).

Germany USA

A B C D

PAH Max Median Max Median Max Mean Max

Naphthalene 4.5 0.20 1.90 – – 0.33 4.3 Acenaphthylene 0.9 0.03 0.10 – – 0.08 0.27 Acenaphthene 1.6 0.05 0.26 – – 0.05 0.18 Fluorene 1.5 0.09 0.24 – – 0.12 1.22 Phenanthrene 1.2 0.96 2.11 – – 0.44 2.15 Anthracene 2.2 0.07 0.21 – – 0.12 0.75 Fluoranthene 22 0.96 3.19 – – 0.52 1.89 Pyrene 14 0.67 2.28 – – 0.43 1.65 Benzo[a]anthracene 24 0.29 1.41 0.136 30.5 0.22 0.69 Chrysene 23 0.55 2.00 0.27 a 28.0 a 0.39 2.41 Benzo[b]fl uoranthene 17 0.54 1.90 0.31 47.0 0.55 b 1.34 b Benzo[k]fl uoranthene 5.5 0.37 1.91 0.099 13.1 Benzo[e]pyrene – 0.4 1.42 – – 0.26 0.75 Benzo[a]pyrene 19 0.27 1.39 0.154 25.5 0.23 0.63 Indeno[123cd]pyrene 9.6 0.33 1.17 0.161 23.4 0.23 0.70 Dibenzo[ah]anthracene 1.4 0.05 0.29 0.036 4.5 0.10 0.41 Benzo[ghi]perylene 6.8 0.35 1.28 – – 0.25 0.61 Coronene – 0.16 0.47 – – 0.13 0.50

a Including iso - chrysene. b Including benzo[k]fl uoranthene. A: collected from all over Germany ( n = 703), P ö hner et al . (1998) , medians for all PAHs are < 0.2 mg kg − 1 . B: from apartment blocks in Berlin, district Steglitz ( n = 61), Fromme et al . (2004a) . C: collected in the Detroit area, Michigan; the entire state of Iowa, Los Angeles County, and the Seattle area, Washington ( n = 605/616), Camann, Colt and Zuniga (2002) . D: from low income homes, North Carolina ( n = 24), Chuang et al . (1999) .

in one or more of the seven size - fractionated samples. The concentrations of nearly all contaminants increased gradually with decreasing particle size, then increased dramatically for the two smallest article sizes (4 - 25 μ m and < 4 μ m). The B2 PAH content in the coarse dust sample was 2.21 mg kg − 1 .

Concentrations of PAHs measured in house dust are in the same order of mag-nitude for American and German homes (see Table 11.9 ). This holds especially for BaP, the most prominent indicator for PAHs. Values are further in accordance with the median for BaP of 0.2 mg kg − 1 given by Kersten und Reich (2003) for house dust from 65 apartments in Hamburg (Germany).

Wilson et al . (2003) , studying the aggregate exposures of preschool children to persistent organic pollutants at day care and at home, reported that the indoor


Table 11.10 Concentrations of selected polycyclic aromatic hydrocarbons ( PAH s ) in indoor air from Germany and the USA (ng m − 3 ).

PAH Germany Fromme et al . (2004b) a

USA Chuang et al . (1999 ) b

Median Max Mean Max

Naphthalene – – 2190 9700 Acenaphthylene – – 43.9 533 Acenaphthene – – 96.0 298 Fluorene – – 49.7 196 Phenanthrene – – 66.3 184 Anthracene – – 6.26 16.7 Fluoranthene 0.34 1.86 8.58 19.3 Pyrene 0.19 4.57 6.73 29.4 Benzo[a]anthracene 0.04 0.67 0.59 3.13 Chrysene 0.19 1.42 0.84 3.38 Benzo[b]fl uoranthene 0.10 1.98 1.27 c 5.80 Benzo[k]fl uoranthene 0.05 0.76 Benzo[e]pyrene 0.33 2.25 1.05 5.28 Benzo[a]pyrene 0.09 1.22 0.70 4.49 Indeno[123cd]pyrene 0.24 1.71 0.88 8.80 Dibenzo[ah]anthracene 0.07 0.28 0.54 2.75 Benzo[ghi]perylene 0.19 1.50 1.46 20.9 Coronene 0.14 3.35 1.15 17.1

a From apartments in Berlin, sampled in 2000, non - smokers ( n = 61). b From low income homes, North Carolina, sampled in 1994/1995, smokers and non - smokers

( n = 24). c Including benzo[k]fl uoranthene.

exposure to PAHs (like to other SVOCs), was greater than the outdoor exposure. Uptake of PAHs was of similar magnitude for the day care centers and the homes.

No association was found for BaP in house dust and the indoor air by Dieckow, Ullrich and Seifert (1999) . Nor did Fromme et al . (2004b) fi nd an association between concentrations of polycyclic aromatic hydrocarbons in household dust and indoor air. But for fl uoranthene, phenanthrene, and pyrene, statistically sig-nifi cant correlations of surface dust loadings (ng cm − 2 ) with personal air and indoor air concentrations were noted (Clayton et al ., 2003 ). As stated by Hansen and Volland (2002) , there seems to be no simple relation between PAHs in air and in dust.

Fromme et al . (2004b) summarized the factors infl uencing PAHs in indoor air and dust. These include open fi replaces, domestic heating with coal burning stoves and smoking or incense burning. Concentrations especially in indoor air are infl uenced by traffi c volume (Fromme et al ., 2004b ) but most signifi cantly by smoking (increase by a factor of 59). Concentrations in house dust are especially


infl uenced by the presence of tar and bitumen based adhesives. BaP concentra-tions in house where these adhesives have been used may amount to some grams per kilogram (Dieckow, Ullrich and Seifert, 1999 ).

11.3 Sources for SVOC s Indoors

Sources for contaminants indoors are: (i) the indoor environment (ii) compounds tracked in from outdoors and (iii) indoor applications especially of biocides (pes-ticides). Indoor sources include permethrin from impregnated carpets and textiles, wood preservatives like pentachlorophenol, lindane and dichlofl uanid from treated timber, PAHs from glues of parquet fl oors, phthalates and PCBs from plastics or sealing material and polychlorinated diphenyl ethers from fl ame retardants used in electronic equipment. In many cases consumers are not aware of the fact that upon buying equipment and furnishing, or upon renovating or remodeling their homes, they bring chemicals into their homes which might disperse in the indoor environment.

Household attributes may infl uence the biocide content especially in house dust. Signifi cantly higher concentrations of permethrin were measured for housings with wool carpets and wall - to - wall carpeting compared with those not bearing this attribute (Butte, 2003 ). For both BBzP and DEHP, associations were found between house dust concentrations and the amount of PVC used as fl ooring and wall mate-rial in the home. Furthermore high concentrations of DEHP were associated with buildings constructed before 1960. But BBzP and DEHP were found in buildings with neither PVC fl ooring nor wall covering as well (Bornehag et al ., 2005 ).

Biocides are known to be tracked into the homes after a certain outdoor applica-tion (Lewis and Nishioka, 1999 ; Nishioka et al ., 2001 ). Furthermore transport routes, that is, transfer from the workplace to the home (para - occupational or take - home exposure) may be relevant. Household dust concentrations of azino-phosmethyl, chlorpyrifos, parathion and phosmet were signifi cantly lower in refer-ence homes when compared with farmer/farmworker homes (Simcox et al ., 1995 ). Dialkyl phosphate metabolites measured in children ’ s urine were elevated for agrarian children compared with children whose parents did not work in agricul-ture (Fenske et al ., 2000 ).

Regarding indoor applications of biocides, higher PCP concentrations in house dust of homes where wood preservatives had been used compared with those without an application were reported (Walker, Hostrup and Butte, 1999 ; Schulz and Butte, 2007 ). DDT, methoxychlor and PBO concentrations in dust of homes where insecticides had been used ( ‘ users ’ ) were signifi cantly higher than in those of non - users; occupants stating they fought insects revealed higher concentrations of propoxur in house dust (Walker, Hostrup and Butte, 1999 ). When comparing biocide levels in carpet dust to self - reported pest treatment practices in four USA sites, Los Angeles residents reported the most treatment for crawling insects, fl eas/ticks, and termites, and Los Angeles dust samples had the highest levels of

11.4 The Indoor Environment: A Source for Exposure? 261

propoxur, chlorpyrifos, diazinon, permethrin and chlordane. Iowa residents on the other hand had the most treatment for lawn/garden weeds, and had the highest levels of 2.4 - D ( 2.4 - dichlorophenoxy acetic acid ) and dicamba (Colt et al ., 2004 ).

11.4 The Indoor Environment: A Source for Exposure?

Children are the most susceptible population to contaminants indoors, especially in house dust because: (i) their organs and neurological system are rapidly devel-oping, (ii) their intake is greater relative to body size and weight, (iii) their activities on and proximity to the surfaces enhance their potential contact with toxic sub-stances (IPCS, 1986 ).

11.4.1 Indoor Air and House Dust: Associations to Human Biomonitoring

For some biocides indoor air may be the most important route of incorporation. Pang et al . (2002) estimated that inhalation of indoor air accounted on average for 84.7% of the aggregate daily exposure to chlorpyrifos with exposure rates that were signifi cantly correlated to concentrations in indoor air and carpet dust. Inhalation exposure also exceeded dietary exposure for cyclodiene insecticides like aldrin and chlordane and for pesticides used mainly in the home in a study designed to assess total human exposures to 32 pesticides and pesticide degradation products in the non - occupational environment (Whitmore et al ., 1994 ). A signifi cant correlation was observed between personal air concentrations of chlorpyrifos, diazinon, and propoxur and concentrations of these insecticides or their metabolites in plasma samples (maternal and/or cord) (Whyatt et al ., 2003 ). Exposure of pupils to PCBs in indoor air of contaminated schools caused increased blood concentrations of the lower chlorinated congeners. However, compared with background levels, the detected excess body burden was very low indicating no additional health risk and exposure was not associated to any specifi c subjective complaints (Liebl et al ., 2004 ).

Regarding house dust, associations between SVOC contaminants in house dust and in blood or urine are contradictory. Mei ß ner and Schweinsberg (1996) reported an association between PCP concentrations in passively deposited particulate matter and the concentration in urine. A signifi cant correlation for PCP in house dust and urine was also found by Krause and Englert (1980) and Schulz and Butte (2007) for PCP in house dust and concentrations in plasma, respectively (Butte and Heinzow, 1998 ). But results from Liebl et al . (1996) and Rehwagen et al . (1999) did not confi rm the association between PCP in urine and house dust.

Becker et al . (2004a) , analyzing the levels of the urinary DEHP metabolites and DEHP in house dust in a study including 254 children aged 3 to 14, did not fi nd an association. No association was further observed for concentrations of perme-thrin in house dust compared with permethrin metabolites in urine by Butte,


Walker and Heinzow (1998) , but Becker et al . (2004b) reported that permethrin in house dust was one of the signifi cant predictors of pyrethroid metabolites in the urine of children.

The contradictory results may be due to different contamination levels. With a high indoor contamination it might be easier to detect an association, whereas with low indoor contamination levels associations may be hidden by the ubiqui-tous presence of the biocide and dietary intake. Taking DDT as an example, that formed much higher residues in house dust 30 years ago, it was stated in 1975 that the contamination of house dust by the domestic use of DDT was primarily responsible for human serum residues (Davies, Edmundson and Raffonelli, 1975 ).

Also for herbicides like 2.4 - D re - suspended fl oor dust may be a major source of exposure. Estimated post - application indoor exposure levels for young children from non - dietary ingestion were calculated to be 1 – 10 μ g per day from house dust in contact with fl oors, thus being about eight times higher than the dietary ingestion of approximately 1.3 μ g per day (Nishioka et al ., 2001 ). In a review of Lioy, Freeman and Millette (2002) house dust and residential soil were examined in detail regarding their use for identifying sources and for estimating levels of exposure.

11.4.2 Indoor Biocides: A Reason for Health Impairments?

Since about 10 years a possible association between health impairments (especially childhood cancer) and an indoor exposure with biocides is discussed, but biocide concentrations in the indoor environment were seldom measured in detail (Daniels, Olshan and Savitz, 1997 ). Information about frequency and mode of biocide use were mostly obtained through questionnaires instead (e.g., Buckley et al ., 2000 ; Infante - Rivard et al ., 1999 ; Kato et al ., 2004 ; Leiss and Savitz, 1995 ; Meinert et al ., 2000 ; Menegaux et al ., 2006 ).

Nowadays evidence is growing that indoor pesticide exposure is of particular concern for pregnant women (Berkowitz et al ., 2003 ). Results indicate that prenatal exposures to chlorpyrifos in indoor air may have resulted in impaired fetal growth and that diazinon exposures may have contributed to the effects (Whyatt et al ., 2004 ). Ma et al . (2002) reported that the use of professional pest control services at any time from 1 year before birth to 3 years after was associated with a signifi -cantly increased risk of childhood leukemia. Insecticide exposures early in life appeared to be more signifi cant than later exposures, and the highest risk was observed for exposure during pregnancy. Additionally, more frequent exposure to insecticides was associated with a higher risk. In contrast to insecticides, the association between herbicides and leukemia was weak and non - signifi cant. Hartge et al . (2005) as well found no increase in risk of non - Hodgkin - lymphoma ( NHL ) for herbicide use on the lawn or in the garden. Even residents who had the highest dust levels had no elevated risk.

In contrast to herbicides the use of insecticides seems to be associated to an elevated risk for leukemia and NHL for adults as well as for children. A study performed in Germany, that is, the ‘ Study on Leukemia and Malignant Lymphoma

11.4 The Indoor Environment: A Source for Exposure? 263

in Northern Germany (NLL) ’ underlined the hypothesis, that exposure to pesti-cides (i.e., insecticides and fungicides) in the indoor environment increases the risk for leukemia and malignant lymphoma (Hoffmann et al ., 2008 ). Colt et al . (2005) reported the NHL risk to be elevated if any of the PCB congeners (PCB 105, 138, 153, 170 or 180) was detected in carpet dust that served as an exposure indica-tor. NHL risk was also elevated if DDE was detected, but only among men. In a study published recently Colt et al . (2006) further found a signifi cant trend of increasing risk for the linkage of NHL to exposure with α - Chlordane. Levels of insecticides in dust taken from used vacuum cleaner bags were used to quantify the exposure.

11.4.3 Reference and Guideline Values

For the interpretation of monitoring results in general two approaches might be used: (i) the comparison with a reference values obtained from representative studies and (ii) a risk assessment methodology related to the hazard of the com-pound. The latter results in guideline values.

Reference values are intended to characterize the upper margin of a current background contamination of a pollutant at a given time. For contaminants in the indoor environment reference values may be calculated applying the procedure as for environmental toxins in body fl uids (Ewers et al ., 1999 ). Reference values for pollutants in house dust were reviewed recently by Butte (2003) . It has to be emphasized that a reference value has no health meaning nor regulatory implication.

Applying a risk assessment methodology, guideline values may be deduced from LOAELs (lowest observable adverse effect level), see Figure 11.3 . This holds for contaminants in air as air is inhaled and contaminants are expected to be absorbed and incorporated. But guideline values are only available for a few SVOC, that is, pentachlorophenol (Ad - hoc Working Group, 1997 ), tris(2 - chloroethyl) phosphate (Sagunski and Ro ß kamp, 2002 ) and naphthalene (Sagunski and Heger, 2004 ). Furthermore air analysis has a number of shortcomings: valid analysis is time consuming and expensive and concentrations fl uctuate considerably due to varia-tions in room ventilation, humidity and temperature.

Guideline values for SVOCs in house dust are not available as yet. House dust has received only limited attention in the past with regard to a risk assessment process, although health effects have been closely associated with indoor biocide applications and biocides in house dust (see Section 11.4.2 ). Small children are considered to be the population at highest risk since they spend most of their time indoors and much of this time is spent in contact with fl oors, engaging in mouth-ing of hands, toys and other objects (Lewis and Nishioka, 1999 ). Butte and Heinzow (2002) estimated the risk associated with the ingestion of contaminated dust exemplarily by using the chronic oral reference dose ( RfD ) available from the US - EPA Integrated Risk Assessment Information Service ( IRIS , US - EPA, 2006 ). They focused on small children (age 1 – 6 years, mean body weight 16 kg) and


Figure 11.3 Deduction of guideline values for TCEP in indoor air.

assumed a daily intake of 100 mg house dust (Lewis, Fortmann and Camann, 1994 ) for the calculation of tentative benchmarks. The assessment indicated that the tolerable exposure concentration for chlorpyrifos, DDT and diazinon in house dust is exceeded in some cases and chlorpyrifos especially has to be considered a potential hazard to householders.

11.5 Summary

SVOCs found in the indoor environment origin form different sources, as dis-cussed in Section 11.3 . According to the sources and uses of SVOCs, concentra-tions differ signifi cantly, that is, more than a factor of a billion (see Figure 11.4 for some selected compounds). Concentrations range from gram per kilogram (phthalates like DEHP), down to some nanograms per kilogram for 2.3.7.8 - tetrachlorodibenzo dioxin.

Semi - volatile organic compounds in the indoor environment are a relevant source of exposure for the occupants, but guideline values for SVOCs in air are

References 265

Figure 11.4 Range of concentrations of contaminants in dust.

rare; they are not available for house dust. An approach based on the reference dose for chronic exposure might be used instead temporarily.

Indoor air is an ideal material to get information about the actual exposure via inhalation. House dust, on the other hand, as a measure of indoor contamination, is easily accessible, compounds are stable, shipment is simple and concentrations are high compared with air.

Advances in construction technology, cost reduction in building furniture, new electronic equipment etc. led to a greater use of synthetic materials and chemicals. Furthermore changes in building design have made modern homes more airtight than older structures. This led in principle to more comfortable buildings with lower running costs, but as a consequence semi - volatile organic contaminants originating from different sources may build up indoors to much higher concen-trations than before.

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Indoor Pollutants in the Museum Environment Alexandra Schieweck , Tunga Salthammer , and Simon F. Watts

12

12.1 The Museum Environment: An Introduction

In contrast to interiors used for habitation and work, museums, archives and gal-leries have a special position in the context of indoor air pollution. Besides provid-ing a healthy indoor climate for museum staff and visitors, closely defi ned climatic requirements are often necessary to protect the works of art inside. Primary responsibilities for cultural heritage institutions are to collect, to investigate and to preserve cultural assets as well as to hand them on to future generations. There-fore, the topic of preventive conservation is an important part of restoration work. It is defi ned as effort to decelerate the deterioration of cultural heritage, to preserve its integrity and to reduce the necessity of restoration treatment to a minimum (Charter of Vantaa, 2000 ). In recent years, the topic of indoor air quality in museum institutions has become a major concern within preventive conservation. Since the dawn of the industrial age, the effects of air pollutants both on monu-ments located outside and on artifacts stored inside buildings have become obvious, sometimes in very short time intervals. Impacts on painting surfaces caused by the notorious London smog were observed in the National Gallery London, which was located close to various coal - burning industries and their large chimneys: dust deposited, lead white darkened and copper corroded as a result of sulfurous gases. Attempts to minimize them are described in detail by Saunders (2000) . Whereas this report refers to bad indoor air quality due to exogenous pol-lutants and damages resulting from inappropriate climatic conditions, one of the fi rst documentations regarding adverse effects on objects due to pollutants gener-ated indoors is bequeathed from the end of the nineteenth century. Byne (1899) documented effl orescence on sea shells, as discussed in Section 12.4 . However, the main concern of conservation scientists was on climatic infl uences through fl uctuations in temperature and relative humidity ( RH ) as well as on light intensi-ties (Padfi eld, 1966 ; Michalski, 1993 ; Saunders and Kirby, 1996 ; Camuffo, 1998 ; Brimblecombe et al. , 1999 ; Camuffo et al. , 1999 ). The fi rst scientifi c basis of indoor air quality in the museum fi eld was set out by Thomson (1965) , who gave a basic


273

274 12 Indoor Pollutants in the Museum Environment

review of atmospheric chemistry for conservators and dealt with the fundamental parameters affecting the museum environment. In 1978, the fi rst edition of his popular book The museum environment was published (Thomson, 1986 ). At nearly the same time, Hackney (1984) gave the fi rst comprehensive survey of the distribu-tion of gaseous air pollutants in museum interiors. Consequently, indoor air quality gained growing awareness in the restoration community. The emphasis of subsequent surveys was dedicated to (i) occurrence and concentrations of air pollutants in museum interiors (Grzywacz and Tennent, 1994 ; Oreszczyn, Cassar and Fernandez, 1994 ; Bradley and Thickett, 1999 ; Cassar, Blades and Oreszczyn, 1999 ; Hatchfi eld, 2002 ); (ii) possible emission sources (Brimblecombe, Shooter and Kaur, 1992 ); and (iii) their potential to cause adverse effects on artifacts (Bro-kerhof and van Bommel, 1996 ; Dupont and T é treault, 2000 ). The focus was pri-marily on inorganic compounds such as ozone (O 3 ), nitrogen oxides (NO x ), sulfurous gases (SO 2 , H 2 S, OCS) as well as on organic acids (formic acid and acetic acid) and formaldehyde (Whitmore and Cass, 1988, 1989 ; Williams, Grosjean and Grosjean, 1992 ; T é treault and Stamatopoulou, 1997 ; T é treault, Sirois and Stamato-poulou, 1998 ; Raychaudhuri and Brimblecombe, 2000 ; Ankersmit, Tennent and Watts, 2005 ). Initial investigations of levels and effects of VOCs and SVOCs, including the contamination of museum collections with biocides, have also been carried out recently (Glastrup, 1987 ; Kroo ß and Stolz, 1993 ; Leimbrock and Wagner, 1998 ; Unger, 1998 ; Unger, Schniewind and Unger, 2001 ; Schieweck et al. , 2005, 2007a ).

Inorganic and organic substances pass into the building with the external air through windows, doors and structural leakages, but they can also be generated inside by a variety of potential emission sources. Therefore, in regard to museum institutions, to the three classes of general emission sources of indoor air pollut-ants ( Moriske, 2000 ):

• environmental effects (external air, ground); • human activities (breathing, transpiration); • building products, furniture and fi xtures;

we can add a fourth:

• exhibits (materials and products used for conservation and/or restoration purposes).

Therefore, not only packaging and building products, but also the artifacts themselves contribute to the indoor air pollution by generating different sub-stances causing adverse effects both on human health and on cultural assets (Schieweck et al. , 2005 ). Infl uences and interactions in the museum environment are described in Figure 12.1 .

In order to achieve a comprehensive control of the surrounding conditions, many museum institutions today use progressive containment as an underlying strategy to preserve artifacts in an ideal way. In order to minimize environmental infl uences (climate fl uctuations, entry of polluted outer air), climatic requirements

12.1 The Museum Environment: An Introduction 275

within (i) the museum building and (ii) inside galleries, exhibition areas and repositories are closely defi ned. Climatic conditions within galleries are more controllable than those in the building as a whole.

Furthermore, to prevent mechanical damage (theft, attacks) and to ensure a microclimate inside, which is independent of the surrounding room to suit the individual requirements of the specifi c artifact, mobile cultural assets are stored and displayed in showcases, cabinets and envelopes; paintings are glazed. This kind of nested prevention strategy is referred to as a so - called ‘ box in a box - model ’ (Camuffo, Sturaro and Valentino, 2000 ). It involves two related assumptions: (i) the outside atmosphere is harmful to the artifact, and (ii) increasing benefi t is afforded by increasing degrees of containment and control. While an artifact may gain protection from people and the external climate, this approach can neverthe-less be deeply fl awed. This is because it neglects the chemical interactions between the object on display and the case atmosphere. The assumption that the outside air is always more harmful to the artifact than the air from inside the museum is often not supported by the facts (Brimblecombe, Shooter and Kaur, 1992 ), and that the case atmosphere is the most controlled and most benefi cial of all, is in some situations plainly wrong (Sease et al. , 1997 ). Many products traditionally

Figure 12.1 Outdoor and indoor emission sources affecting human health and works of art in the museum environment (reprinted from Schieweck et al. , 2005 with permission from Elsevier) .


used for display and storage must be considered as sources of hazardous com-pounds which can harm works of art. Especially formaldehyde and organic acids (formic acid, acetic acid) were targeted in the past due to their corrosion potential. Thus, there are efforts to avoid and to replace materials, which are known to emit corrosive compounds, by low - emission fabrics or to improve the conditions by coatings and absorbers (Miles, 1986 ; Grosjean and Parmar, 1991 ; T é treault and Stamatopoulou, 1997 ; Hatchfi eld, 2002 ).

In order to meet the above - mentioned conditions and also to minimize dust entry, modern cases are predominantly highly sealed, which implies that the air exchange rate would be lowered to a minimum. It is to be assumed that the high surface to volume ratio, which especially characterizes smaller enclosures, com-bined with unsuitable construction materials and almost static conditions, enhance the accumulation of chemical compounds. Hence, the comparison of the indoor environment as a ‘ reaction vessel ’ , as stated by Weschler and Shields (1997) , comes to a head in museum showcases.

This sets the context for examining air quality in museums; currently important discussions hinge around the identifi cation of frameworks, defi nitions and the forms of guidelines or standards for indoor air quality from the perspective of cultural heritage (T é treault, 2003 ; Grzywacz, 2006 ). These will be discussed in Section 12.9.1 .

This chapter summarizes the current knowledge, characteristics and some of the more important reaction mechanisms underlying indoor air quality. Effects of VOCs and SVOCs on human health are reported elsewhere and will not be repeated here.

12.2 Climatic Conditions

The lifetime of an artwork is signifi cantly dictated by environmental factors: in particular RH and temperature as well as their rate of change (temporal varia-tions). An artifact must be viewed as an object which is being conserved usually long beyond the design lifetime envisaged by its makers. This means that when an object begins to show signs of damage (whether chemical or physical) it is often underlain by structural damage caused by the effects of climatic factors over extended periods of time. At the molecular level, solid objects contain structural defects; these defects usually have higher energy than the nondefective parts of the structure, which means that further corrosion or damage will take place at the defect sites. Effectively, the object has a memory, and becomes progressively more vulnerable. Thus, it is often the case that this structural damage increases the rate of subsequent chemical damage for thermodynamic reasons. Of course the damage can also be caused by an inappropriate environment in which the object is placed and which is markedly different from the surrounding conditions that had been prevalent at the time of its manufacture. Looking at the two major vari-ables (temperature and RH) separately:

12.2 Climatic Conditions 277

12.2.1 Humidity

The desired or guideline values for humidity in museums are usually predicated on the management plans for different types of artifact (Waller, 1994 ; Ashley - Smith, 1999 ), and expressed in a long term target RH value and an allowed devia-tion from this value. For galleries, the target values vary internationally depending on the climates of those countries (Padfi eld, 1994 ; T é treault, 2003 ), see Table 12.1 .

There are two main effects of humidity on artifacts: enhanced rate of chemical processes and structural damage. All materials may be split into two broad classes: those which absorb water vapor (e.g., unglazed ceramics, wood, leather etc.), termed absorptive; and those in which changes in humidity only result in changes on the thickness of the water layer on the surface of the exhibit (at all reasonable humidities), termed refractive. There has been much work on the response of absorptive artifacts to humidity change, and recently the focus has moved to dimensional change (Camuffo and Pagan, 2004 ). Because the timescales of mois-ture absorption are often longer than that of the changes themselves, such materi-als often act as humidity buffers, reducing the measured humidity changes in cases and galleries. However, absorption of water (or loss of water) from an artifact

Table 12.1 A selection of target values for R H for museums and galleries, including those for particular classes of artifact.

Artifact class Refractive Absorptive

Examples Metals, plastics, ceramics, ivory, glass, magnetic tape

Wood, cellulose nitrate, bone, textiles, leathers, paper, stone, cellulose acetate, paintings, photographic fi lm

Target Values (1) 50 50

Target Values (2) 55 a 55 a

Target Values (3) < 65 a < 65 a

Target Values (4) 45 – 55 a , b 45 – 55 a , b

Target Values (4) 45 ± 10 a , c 45 ± 10 a

Target Values (5) Locally contingent Locally contingent

Notes: (1) T é treault (2003) Average humidity at which LOAED gives CP (conservation period) > 5 years for most sensitive representative at ambient pollutant level; (2) Thomson (1986) ; (3) USA (Northern Collections Centre) specifi es rates of humidity change for different classes of artifact; (4) Michalski (2000) ; (5) Padfi eld (1994) . a with temperature control. b for records required to maintain strict dimensional stability. c Mixed collection – low temperatures, low humidities and stability of humidity.


often results in structural changes (Camuffo and Pagan, 2004 ), particularly in the object ’ s volume. This leads to structural stresses and often damage to the artifacts. Case studies are plentiful, but a particularly good example involves wooden statues in Italian churches (Camuffo and Pagan, 2004 ). A critical factor here is not only the actual humidities involved, but the rate of change between the two humidity states; generally the steeper the rate of change, the less able are the artifacts to accommodate the volume changes without incurring structural damage.

The second main effect of humidity on artifacts is enhanced rates of destructive chemical processes that take place at higher RHs. This most often is shown on refractive artifacts, and is a result of enhanced chemistry occurring as a result of thicker surface water levels as a result of higher RH (Graedel, 1992, 1994 ). A good example is metal degradation; tarnish and corrosion of both silver and copper are enhanced at higher RHs (Graedel et al. , 1985 ). Detailed mechanisms will be dealt with under Section 12.3 .

There are a number of approaches to RH control in buildings, from the use of air conditioning systems through to passive design of buildings (Padfi eld, 1999 ; Blades et al. , 2000 ). It is true to say that most museums use a mixed approach to these issues; over - reliance on air conditioning has been shown to be potentially damaging to artifacts (Grzywacz and Tennent, 1994 ), as can be the laissez - faire approach, although most museums are faced with the challenge of re - using exist-ing buildings with limited resources. Something which increasingly seems to fi nd application is the use of high density new builds (giving temperature and hence RH stability) with light touch air conditioning combined with large scale RH buffering (Padfi eld, 1999 ). This can be expensive in terms of oncost, but has much lower running costs in the long term over large scale ‘ technical fi xes ’ . Good examples might be the Elizabeth Fry building (University of East Anglia, UK) or the new Pitt - Rivers Museum building (Oxford, UK).

12.2.2 Temperature

There are two main effects of temperature: the fi rst is that temperature affects and effects the rate of any chemical reaction, so the higher the temperature, the higher the rate of reaction (e.g., a tarnish or damage reaction); the second is that tempera-ture directly affects the RH because air can hold different amounts of water vapor at different temperatures. Very much a prerequisite of reliable RH control is good temperature control. It is also often the case that historically, guideline values in many museums have tended to be predicated on human comfort.

12.3 Inorganic Atmospheric Compounds

Atmospheric components get into the indoor atmosphere by one of three routes:

12.3 Inorganic Atmospheric Compounds 279

• They are normal compounds of the external atmosphere, which get into the interior by exchange with the outside air through ventilation, bricking joints or leaky windows.

• They are produced chemically in the indoor atmosphere.

• They are added into the atmosphere by emissions of the furniture, fi xings, the building itself or by the artifacts themselves (human emissions are dealt with separately below; see Section 12.8 ).

There are probably two perspectives through which indoor air pollution can be viewed, they are:

1. Generally the indoor environment allows different chemical transformation reactions to occur than usually predominate in the outside atmosphere. So called ‘ night - time chemistry ’ (atmospheric chemical reactions not driven by photochemistry) is usually a good starting point to consider the in - museum chemistry that goes on.

2. Indoor environments usually have lower concentrations of the ‘ outdoor ’ pollutants than the outdoor atmosphere as many museums are equipped with air conditioning and various fi ltration systems, and of course additionally indoor surfaces (including artifacts) can act as reaction sites and sinks of pollutants by virtue of reaction and adsorption processes.

The eventual indoor atmosphere will be the result of the two processes above acting on an atmosphere which is augmented by emissions from the building materials, contents and people in the building. For example, with nitrogen species it has been shown that relatively innocuous nitrogen compounds (e.g., N 2 O) with the help of standard Plaster of Paris, and night - time chemistry can produce atmo-spheric nitric acid (HNO 3 ) and nitrogen dioxide (NO 2 ) (Carslaw, 2003 ).

An on - line model has been developed for conservators to help them estimate indoor atmospheric concentrations of the common museum pollutants (Blades et al. , 2000 ). This is especially useful if used in conjunction with targeted analytical campaigns to assess the atmosphere in a museum or gallery, although the inter-pretation of the results of such measurements is usually not straightforward (Section 12.8 ).

Beyond the composition of the indoor atmosphere itself, the interaction of the atmosphere with not only the building materials (above) but also the artifacts must be considered. Some classes of damage are understood at one level, but in reality as each artifact is a ‘ one off ’ ; the degradation often presents in different ways even though it may be underlain by similar general chemical models, for example, ‘ Bynes Disease ’ discussed below. The tarnishing of metals (Pb, Ag and Cu) is a good example. All three metals are tarnished by reduced sulfur gases, such as hydrogen sulphide (H 2 S) and carbonyl sulfi de (OCS), and the rate of tarnish is increased by humidity, but the way the tarnish presents under different circum-stances varies enormously from fi lamentous to general tarnish. The reactivity of


the artifact is also a major factor in how fast an object tarnishes (Ankersmit, Tennent and Watts, 2005 ).

Although the exact mechanism(s) of the surface tarnishing process with hydro-gen sulfi de on these metals is not fully understood, for silver the fi rst stage seems to be an adsorption into the surface water layer, followed by an electrochemical dissolution process (Graedel et al. , 1985 ; Graedel, 1992 ). Schematically, the systems may be represented as:

H S H O HS H O2 2 3+ → +− +

21

22 2Ag HS Ag S H+ → + +− −e

Where OCS is the tarnishing agent, for either metal it has been shown that the fi rst stage is the adsorption of the corrodent into the surface water layer (Graedel et al. , 1985 ; Graedel, 1992 ), analogous to the tarnishing of silver with H 2 S.

In contrast to the case for copper, for silver a reduction of the tarnish rate by circa 100 - fold has been observed as tarnish layers increase (Graedel et al. , 1985 ), thus indicating that to a degree, the tarnish layers themselves impede further tarnishing. It is known that major silver tarnish can occur at concentrations of circa 100 ppt v and it has been shown that in effect it is the sum of OCS and H 2 S along with artifact reactivity, which determines the tarnish rate (Ankersmit, Tennent and Watts, 2005 ).

Much of the discussion above for silver is pertinent for copper. Although the general mechanism of fi lm growth seems similar, it is thought that the fi rst step is chemical and involves the dissociation of H 2 S on the surface, followed by a redox reaction of sulfur with copper (Fiaud and Guinement, 1985 ):

H S H O HS H O2 2 3+ → +− +

Cu HS CuS H+ → + +− −1

22 e

It is important to note that there is no evidence to indicate that the tarnish layers themselves impede corrosion; this may well indicate that this is a multistep oxidation process involving variable stoichiometry series of semiconducting sulfi des.

There seems to have been relatively little work on lead corrosion, and much of what follows is conjecture. However, it is known that lead tarnish (like Cu and Ag) is dependent on humidity (Graedel, 1994 ), and this implies a role for the surface water layer. In general a PbO passivating oxide layer forms (Grauer and Wiedmer, 1971 ), and the equilibrium dissolution of this produces Pb 2+ , which can react with the dissociation products of dissolved sulfur species (e.g., H 2 S and OCS) to form PbS. Thermodynamically, it appears that although PbS should form fairly readily, it is possible that the kinetic barrier between it and PbSO 4 is considerably lower than that for Ag. There is evidence to support this from studies of Pb tarnish layers which seem to contain little PbS (Graedel, 1994 ). In addition strong oxidiz-ing agents have been observed as a product of Pb corrosion (Dunstan, Jowett and

12.4 Formaldehyde, Organic Acids (Formic Acid, Acetic Acid) 281

Goulding, 1905 ; Thibeau et al. , 1980 ), which also might explain the lack of observed PbS. A qualitative reaction scheme has been proposed (Graedel, 1994 ), but most of the research work remains to be done for this metal.

2 22Pb O PbO+ →

PbO H O Pb H O+ → + ++ −2

22 2 2e

H S H O HS H O2 2 3+ → +− +

HS Pb H O PbS H O− + ++ + → +22 3

PbS O H O H O PbSO H O+ + + → + ++ −2 2 2 2 4 32 2 2e

12.4 Formaldehyde, Organic Acids (Formic Acid, Acetic Acid)

Organic compounds containing carbonyl groups (C = O) released from materials used for furnishing exhibition areas and repositories, as well as for producing museum enclosures, are the prime reason for deterioration of museum artifacts. Due to their well - known corrosive potential, formaldehyde and organic acids (formic acid and acetic acid) are among the most discussed and investigated sub-stances in the museum environment. Both formaldehyde and formic acid and acetic acid, respectively, are categorized into the group of VVOCs with boiling points ranging from < 0 ° C to 100 ° C.

Formaldehyde is the most investigated chemical pollutant, presumably due to its ubiquitous existence in the indoor environment (Finlayson - Pitts and Pitts, 2000 ) resulting from its widespread industrial use, particularly as a component of adhesive agents (urea - formaldehyde, phenol - formaldehyde and melamine - formal-dehyde resins) in wood - based materials. The widely known potential of these resins to emit formaldehyde results either from hydrolysis of the cured resin or from the emission of residual formaldehyde due to incomplete polymerization during the production process. Consequently, great efforts have been made to reduce these formaldehyde emissions by reduction of free formaldehyde or by substitution of formaldehyde by other substances, but nevertheless there are in either case residual concentrations. This also applies to natural wood, whose formaldehyde emissions are due to the decomposition of lignin under the infl u-ence of light (Meyer and Boehme, 1997 ). Also organic acids (formic acid and acetic acid) are always detectable in indoor air. Acid emissions from wood and wood - based products result from the hydrolysis of acetyl groups of hemicelluloses and of side - chains of lignin, respectively (Fengel and Wegener, 1989 ). Besides this, a large range of indoor products release acetic acid vapor: paints, coatings, adhe-sives, for example, products based on polyvinyl acetate, acid - cured silicone and household cleaning agents.

However, the primary emission sources of these compounds in the museum environment are – besides acid - curing sealants and cleaning agents – without con-


troversy wood and wood - based products. They are favored for furnishing storage rooms and display cases as well as for constructing platforms and movable walls due their availability, low - cost, practicability and hygroscopicity. Especially in closed boxes such as display and storage cases, cabinets and drawers, the risk of formaldehyde and acid emissions becomes most extreme due to the almost static conditions and the high surface - to - volume ratio. Different pollution monitorings in storage and display cabinets with targeting of carbonyl compounds were accom-plished (Grzywacz and Tennent, 1994 ; Kontozova et al. , 2002 ; Schieweck et al. , 2005 ; Salthammer et al. , 2006 ). Elevated concentrations of formaldehyde and acetic acid were detected in wooden drawer cabinets and wooden display cases, whereas lowest values were analyzed in open stores.

Furthermore it was found that under those nearly static conditions inside museum enclosures, secondary reactions proceed. The main important mecha-nism is ester hydrolysis. Acetyl esters, which constitute a large part of solvents and additives used today for the construction of modern showcase types, are split with the release of acetic acid and the corresponding alcohol. This reaction explains the still elevated acid concentrations inside modern museum showcases, even though no acidic formulations were applied. Based on these fi ndings it became clear that the main emission sources of carboxylic acids have shifted in recent years from wood - based products and acid - curing silicone rubber to solvent borne materials (Schieweck, 2009 ). Unfortunately, the problem of high acetic acid levels have not been solved so far, as a large range of object materials can be affected by carbonyl pollutants: inorganic specimens such as metals, calcareous materials, glass, stone and enamels as well as organic items like cellulose and leather. The corrosion potential of carbonyl compounds increases from formal-dehyde to formic acid and acetic acid. The deterioration process of glass and enamels will be signifi cantly increased by carbonyl pollutants. Sodium methano-ate and sodium ethanoate can be detected as corrosion products (Bradley and Thickett, 1999 ; Torge, Jann and Pilz, 2000 ). Translucent enamels show less stabil-ity against hydrolysis than opaque ones due to their high alkali content and small amounts of chemical byproducts (alkaline earth oxides, aluminum oxide). The susceptibility is also dependent on the mass of coloring oxide in the glass (M ü ller et al. , 2000 ). The most characteristic damage is spicular effl orescence formed as reaction product on the surfaces of metals or calcareous materials and consisting of formates and acetates. Because Byne (1899) was the fi rst to report this phe-nomenon on shells in the end of the nineteenth century, it is called ‘ Byne ’ s disease ’ . He misleadingly supposed that the effl orescence was a kind of contagious disease resulting from remaining soft parts of the shells. The reaction processes of carbonyl acids and calcareous materials are so far not completely understood. Corrosion products on calcareous materials consist in general of hydrated calcium acetate in different forms and appear in combination with chlorides and nitrates (Ryhl - Svendsen, 2001 ). Brokerhof and van Bommel (1996) consider that effl ores-cence on calcareous materials is formed by a direct reaction of acetic acid and an indirect reaction of formaldehyde. The most noticeable circumstance is that

12.4 Formaldehyde, Organic Acids (Formic Acid, Acetic Acid) 283

the effl orescence composition does not inevitably mirror the pollutant composi-tion in the surrounding atmosphere (Grzywacz and Tennent, 1994 ). Even in a single display case it is possible that different corrosion products are formed (Tennent and Baird, 1985 ). Effl orescence on metal items, notably on lead, is formed as lead formate and basic lead carbonate. Both salts are often identifi ed together on the same artifact. Due to its instability, lead acetate is not normally detected as a corrosion product (Grzywacz and Tennent, 1994 ). The susceptibility of metal artifacts depends signifi cantly on its purity degree. Whereas, for example, lead of high purity ( > 98.8%) is very susceptible, the risk of corrosion decreases with traces of tin, zinc, copper, silver or gold as alloying elements (Grzywacz and Tennent, 1994 ; T é treault, Sirois and Stamatopoulou, 1998 ). Similar to other damage on cultural assets induced by indoor air pollutants, the reaction is highly infl uenced by the prevailing climatic conditions, the moisture on the objects surface, the composition of the item as well as by its conservation history (Bradley and Thickett, 1999 ). Salts on the surface of sea shells or archaeological fi nds or salts remaining from previous conservation treatment can initiate and/or catalyze the process and this even at low levels of RH (Tennent and Baird, 1985 ; Brokerhof and van Bommel, 1996 ; Bradley and Thickett, 1999 ; Thickett and Odlyha, 2000 ). Due to the porosity of the corrosion products further diffusion of carbonyl com-pounds to the surface of the item, resulting in an increasing corrosion layer cannot be excluded (T é treault, Sirois and Stamatopoulou, 1998 ), even though this process has not yet fully been explored. Also cellulose materials depolymerize due to acid hydrolysis notably after long term exposure. Molecular chains are cleaved randomly causing a shortening and decreasing strength durability of the fi bers (Dupont and T é treault, 2000 ). If papers are stored in stacks or envelopes, they are more at risk than others kept separately for the following reasons: (i) they are in prolonged contact with pollutants resulting from deterioration pro-cesses from the cellulose itself and (ii) acetic acid diffuses through the whole stack until it reaches equilibrium between the headspace in the box and the stack (Dupont and T é treault, 2000 ).

As mentioned above, the special position of the museum environment is founded by a fourth emission source: the exhibits themselves. In natural history collections formalin, an aqueous formaldehyde solution (30% – 40%), is still used for the conservation of animal specimens as well as formaldehyde and formic acid for the conservation and preparation of zoological exhibits. Schieweck et al. (2005) analyzed formaldehyde concentrations in a zoological collection. The measuring apparatus was located at a distance of 1 m from a metal cupboard in which animal preparations are stored in formalin solutions acting as direct emission source and causing an intense smell. In contrast to normal conditions of use (28 μ g m − 3 ) when the cabinet doors were kept closed, the formaldehyde concentration was increased three times when the doors were opened (90 μ g m − 3 ).

Cellulose acetate products deteriorate by releasing acetic acid. Without air exchange, the acid vapor accumulates and is reabsorbed by the emission source


itself: an autocatalytic reaction proceeds. The typical acid smell has given this reaction the name ‘ vinegar syndrome ’ . It can be observed by shrinking and peeling off the emulsion layer in the case of cellulose acetate fi lms. Some plastics, for example, PMMA or polystyrene, absorb these acid gases and can act as secondary emitters (Fenn, 1995 ). 1)

The role of formaldehyde in harming artifacts has been controversially dis-cussed. Formaldehyde can be oxidized to formic acid in the compartment air or on surfaces via the Cannizzaro reaction. T é treault (2003) has stated that this reac-tion is negligible in the atmosphere without the unusual presence of oxidants, for example, ozone, oxygen, and hydrogen peroxide, and under normal climatic condi-tions. Therefore, the role of formaldehyde has been overvalued in comparison to organic acids. Especially acetic acid is known for its strong impact on a large range of artwork materials, but without showing serious adverse effects on human health (T é treault, 2003 ). The corrosion of lead will be therefore caused either by a direct emission source of formic acid or by formic acid as a reaction product of formal-dehyde (Raychaudhuri and Brimblecombe, 2000 ). Hence, it is not the concentra-tion of formaldehyde in the indoor environment that is important, but rather the probability of its oxidation to formic acid. Because the concentrations of oxidants as well as the levels of RH in typical museum environments are in general clas-sifi ed as low, the oxidation will proceed very slowly (Raychaudhuri and Brimble-combe, 2000 ).

12.5 Volatile Organic Compounds ( VOC s )

Even though VOCs are the most important group of indoor air pollutants due to the variety of their potential emission sources, they have not been investigated with regard to sources and effects in the museum environment up to now. Just a few studies are available so far focussing on the broad spectrum of volatile and semi - volatile organic compounds in storage and exhibition rooms as well as in display cases and diorama (Schieweck et al. , 2005, 2007b, c , Schieweck, 2009 ). Characteristic substances are listed in Table 12.2 . In regard to storage rooms, the varying composition of indoor air pollution refl ects the specifi c furnishing of each museum department, for example, monoterpenes from wood and wood - based materials, alkanes and aldehydes, partially resulting from a recently renovation and DEHP generated by fl ooring materials. The most distinctive difference can be seen in recent decades by comparison of showcases. Whereas in earlier types of cases, wood and wood - based products, and acid - curing silicones were preferred as construction materials, they have been increasingly substituted by so called ‘ inert ’ materials such as metals, powder coatings and neutral - curing adhesives and seal-ants, respectively. Thus, instead of monoterpenes like α - pinene, β - pinene, limo-

1) Original sources not available. Data adapted from Hatchfi eld (2002) .

12.5 Volatile Organic Compounds (VOCs) 285

Table 12.2 Lead compounds of VOC s in indoor air in different locations of museums.

Location Lead compounds

Chemical group Dominating substances

Storage rooms Alcohol 2 - Propanol 2 - Ethyl - 1 - hexanol

Carboxylic acids Acetic acid Aldehydes Benzaldehyde Terpenes α - Pinene

β - Pinene 3 - Carene Limonene Camphor

Aromatic hydrocarbons Chlorobenzenes Naphthalenes

Exhibition rooms Alcohol 2 - Propanol 2 - Ethyl - 1 - hexanol

Carboxylic acids Acetic acid Aldehydes Benzaldehyde Terpenes α - Pinene

β - Pinene 3 - Carene Limonene

Diorama Alcohol Carboxylic acids Acetic acid Aldehydes Hexanal

Benzaldehyde Terpenes α - Pinene

β - Pinene 3 - Carene Limonene

Aromatic hydrocarbons Chlorobenzenes Naphthalenes

Aldehydes Hexanal Nonanal Furfural

‘ Old type ’ showcase Bicyclic monoterpenes α - Pinene β - Pinene Limonene 3 - Carene

Carboxylic acids Acetic acid Aromatic hydrocarbons Toluene

Xylene - isomers Glycol ethers 2 - Butoxyethanol Cyclic siloxanes


Location Lead compounds

Chemical group Dominating substances

‘ New type ’ showcase Alcohol Ethanol Aromatic hydrocarbons Xylene - isomers

Toluene Ethylbenzenes

Carboxylic acids Hexanoic acid Acetic acid

Carboxylic esters Ethylacetate n - Butylacetate

Glycol ethers 2 - Butoxyethanol 1 - Butoxy - 2 - propanol

Glycol esters Ethoxypropylacetate Ethyl - 3 - ethoxypropionate 1 - Methoxy - 2 - propylacetate

Ketoximes 2 - Butanone oxime (MEKO) Methylisobutylketoxime

Cyclic siloxanes


nene, Δ 3 - carene, camphene and others, organic acids such as formic acid and acetic acid and formaldehyde, the dominant substances are currently (di)carboxylic esters and glycol esters, C3 - /C4 - benzenes, aromatic compounds as well as silox-anes and ketoximes. While (di)carboxylic esters, glycol esters, C3 - /C4 - benzenes and aromatic compounds are mainly released by solvent - borne lacquers and coat-ings, siloxanes and ketoximes are generated in increased concentrations by neu-tral - curing silicones used in new showcases as sealants and adhesives, respectively. Among these substances, 2 - butanone oxime (MEKO), a characteristic fragmenta-tion product of neutral - curing silicone rubber, could be detected (Schieweck, Markewitz and Salthammer, 2007b, 2007c , Schieweck 2009 ). However, so far no data are available about adverse effects on cultural assets. Perhaps this lack of knowledge can be attributed to the assumption that high levels of VOCs do not mean a high risk to museum collections because most of these substances are not harmful (T é treault, 2003 ). A certain corrosion potential will merely be acknowl-edged in correlation with potential dangerous compounds in regard to possible synergistic effects (T é treault, 2003 ). Even though adverse effects on artifacts have not been detected so far, further research about interactions of VOCs and artifact substances is necessary. Especially in the case of unhindered exchange between indoor air pollutants and material surfaces, damage cannot be excluded as it is

12.6 Semi-volatile Organic Compounds (SVOCs) 287

well - known that reactive chemicals cause a corrosive environment (Finlayson - Pitts and Pitts, 2000 ).

12.6 Semi - volatile Organic Compounds ( SVOC s )

One observable damage caused by SVOCs is the so called ‘ fogging phenomenon ’ , also well - known under the term ‘ black magic dust ’ (Wensing et al. , 1998 ; Moriske et al. , 2002 ). In the museum environment, fogging can be observed in form of clouded and steamy glasses of display cases, diorama or glazings. These migration processes are attributed to two different mechanisms (i) a special form of co - dif-fusion induced by water clusters and (ii) thermal processes leading to the phe-nomenon of the so called ‘ ghost images ’ , which means that an exact negative of the painting has deposited on the glazing. Both processes are explained on the basis of the ‘ free volume theory ’ (Zumb ü hl, Hons and von Stockhausen, 2004 ). The fi rst process leads to effl orescence of saturated fatty acids in form of thin membranes, which are macroscopically signifi cant due to their high surface area and intensive light dispersion. In regard to the thermal diffusion process (ii), the quantitative extent is selectively dependent on specifi c color areas as they differ in the capability of thermal light absorption and the properties of drying and degrada-tion of oil paints are dependent on the pigment coating. Figure 12.2 shows a chromatogram of a ‘ ghost image ’ . A similar spectrum can also be observed

Figure 12.2 Total ion chromatogram (GC – MS) of a hazy fi lm from a protective glazing of an oil painting (Schieweck and Salthammer, 2006 ).


concerning deposits on other glassy surfaces, for example, trays in display cases. Frequently identifi ed as characteristic substances are long - chained fatty acids and their correlating esters such as tetradecanoic acid, pentadecanoic acid and hexa-decanoic acid. In most cases this phenomenon originates of the exhibits them-selves as migration of object immanent components.

12.7 Occurrence of Biocides in the Museum Environment

Indoor pollution of museum rooms and exhibits due to biocides is a well - known and ubiquitous problem which has been acknowledged for many years (Schieweck et al. , 2007a ). In the effort to preserve the exhibits, different biocides were applied over time. Typical periods of application are shown in Figure 12.3 . Some com-pounds were distributed as pure substances or for free evaporation in cases and diorama. Due to their adverse health effects, biocide active agents pose special problems to museum staff so that extra care must be taken. Notably ethnographic items are highly contaminated, due to their non - European countries of origin and the variability of their organic object materials, for example, leather, feather and cellulose, as well as zoological preparations and wooden objects. In particular, wooden objects were soaked with chemical solutions – especially oily wood preser-

Figure 12.3 Time periods of possible application of biocides in museums (MCN: monochloronaphthalene; DCN: dichloronaphthalene; PCP: pentachlorophenol; DDT: 1,1,1 - trichloro - 2,2 - bis(4 - chlorophenyl) - ethane; 1,4 - DCB: 1,4 - dichlorobenzene) (reprinted from Schieweck et al. , 2007a with permission from Elsevier) .

12.7 Occurrence of Biocides in the Museum Environment 289

vatives – against pests and micro - organism attack during the 1950s and 1960s. Inorganic compounds like arsenic trioxide (white arsenic, As 2 O 3 ), arsenate (V) and arsenate (III) as well as mercury (II) chloride have been used for more than a 1000 years indoors and outdoors for the protection against rodents, insects and micro - organisms. With the beginning of the twentieth century, modern industrial chemi-cal synthesis enabled the extensive distribution of organic agents. Due to ‘ material trends ’ , a variety of chemical products has been used for many years. The most well - known chemical substance is the insecticide 1,1,1 - trichloro - 2,2 - bis(4 - chlorophenyl) - ethane (DDT), which was used worldwide, notably in agriculture and in the abatement of malaria. In conservation it was used since the 1950s and increasingly applied in Eastern Europe and the German Democratic Republic. It was used as the active agent in combination with other compounds under different trade names, for example, Hylotox ® 59 (3.5% DDT, 0.5% lindane) and Hylotox ® IP (3% DDT, 5% PCP). Due to its toxicology it was banned by law in the Federal Republic of Germany in 1972. In the German Democratic Republic the production of Hylotox ® 59 was halted in 1987; however, the use of remaining supplies was permitted until 1990/1991 (Unger, Schniewind and Unger, 2001 ). Besides its high toxicity, artifacts excessively treated with oily wood preservatives containing DDT show blooms due to the migration of the active agent (Unger, 1998 ). Mono - chlorinated naphthalene s ( MCN ) and dichlorinated naphthalene s ( DCN ) were abandoned because of their toxicity and their unpleasant smell and substituted by a combination of pentachlorophenol ( PCP ) and lindane, which was frequently used from the 1950s to the 1970s (Unger, Schniewind and Unger, 2001 ). PCP was banned by law in Germany in 1989.

MCN and DCN were contained as main compounds in several wood preserva-tives, for example, under the trade name Xylamon ® (Germany), Anabol ® (Great Britain) contained chloronaphthalene wax and dichlorobenzenes. These products were introduced into the market in the 1920s (Xylamon ® ) and 1930s (Anabol ® ). Xylamon ® now contains mixtures of lindane and/or permithrine (Unger, Schniewind and Unger, 2001 ). Until the 1970s lindane and pentachlorophenol (PCP) were the main compounds of wood preservatives. PCP is one of the most widespread active agents (e.g., Preventol ® P, Hylotox ® IP) due to its insecticidal and fungicidal properties, and replaced chloronaphthalines (Moriske, 2000 ; Unger, Schniewind and Unger, 2001 ). As a result of its high stability and its suffi ciently high vapor pressure of 2.3 10 − 2 Pa (20 ° C), PCP can be released from materials decades after application and absorb on interior dust and surfaces. Chlorodioxins and furans were detected in rooms in which wood preservatives containing PCP had been applied. PCP was outlawed in Germany in 1989. Due to its adverse health effects the use of lindane is decreasing; the substance is increasingly substituted by other active agents, for example, synthetic pyrethroids, which are still in use (Moriske, 2000 ; Unger, Schniewind and Unger, 2001 ). Whereas compounds such as dichlorobenzenes, lindane, camphor and naphthalenes are detectable in indoor air, inorganic substances like arsenic, chloride and lead, as well as DDT and PCP can mainly be analyzed in dust and wiping samples. Table 12.3 summarizes results of indoor air and dust analyses in different German museums. In contrast


Table 12.3 Dust concentrations of PCP , lindane, DDT and chlorinated naphthalenes in different locations of museums (reprinted from Schieweck et al. , 2007a with permission from Elsevier) .

Museum Location Source Dust type

MCN DCN PCP Lindane DDT Ref.

mg kg − 1 dust

M1 Exhibition room

Sculpture Old < 1 < 1 117 14 n.d. Schieweck et al. (2007a)

M1 Exhibition room

Floor 1 Fresh < 1 < 1 30 5 n.d. Schieweck et al. (2007a)

M1 Exhibition room


M1 Exhibition room


M1 Storage room Cabinet 1 Old 10 – 100 a 10 – 100 a 10 – 50 a 10 – 50 a n.d. Schieweck et al. (2007a)



M3 Depot (South Pacifi c)

Floor Old < 5 < 5 400 26 5 Kroo ß and Stolz (1993)

M3 Reading room Floor Old < 5 < 5 300 22 < 1 Kroo ß and Stolz (1993)

M3 Other rooms Floor Old < 5 < 5 4 – 25 3 – 128 < 1 – 19 Kroo ß and Stolz (1993)

a Range of concentrations; in each cabinet 5 dust samples were taken from different locations: n.d. = not determined.

to the building sector, where the application of biocides is often well documented, the treatment of works of art with wood preservatives and other biocides occurred without record in many cases and is diffi cult to reconstruct (detailed documenta-tion of conservation and restoration work only became an established procedure after the 1970s).

Biocides and pesticides in indoor air are connected with different damage to objects. However, for scientists and museum professionals it is diffi cult to distin-guish damages related to biocide treatment from damages caused otherwise.

12.8 The Role of People 291

Hatchfi eld (2002) gives a review of relevant studies (Jedrzejewska, 1971 ; Tilbrooke, 1975 ; Fenn, 1995 ). 1) After Tilbrooke (1975) , 1) naphthalene is suspected of softening natural resins and of accelerating the corrosion of some metals. 1,4 - dichloroben-zene can cause fading of ink and yellowing of paper (Jedrzejewska, 1971 ). 1) Fur-thermore, it can damage leather or feathers (Dawson, 1984 ). PVC, which was assumed to be stable on exposure to individual biocides, was found to be damaged by exposure to a mixture of biocides (Fenn, 1995 ). 1) Cellulose nitrate and polyvinyl acetate soften, shrink and gain weight by absorbing 1,4 - dichlorobenzene. The same applies to hide glue, which can therefore be a high risk for wooden objects (Dawson, 1984 ). Some polymeric materials such as methyl methacrylate, polycar-bonate and acrylic sheeting, can act as secondary emitters of substances such as acetic acid and biocides as stated in Fenn (1995) . 1)

Today, a focal point of research is on the decontamination of artifacts, especially by means of super fl uid extraction ( SFE ) with carbon dioxide (Unger, 1998 ; Winkler, F ö ckel and Unger, 2002 ; Tello et al. , 2005 ).

12.8 The Role of People

Part of the mission of most museums and galleries will involve the display of artifacts to either the general public or specialist audiences. This in turn means that there can be a confl ict between two of the major missions for the museum ’ s existence: the conservation management of the artifacts on the one hand and the demands to display those artifacts on the other. People can affect indoor atmo-spheres in two ways, either directly or indirectly.

Direct effects. This includes those emissions that come directly from the human body and include water, heat and chemical substances (here we will consider chemical substances to mean gaseous emissions only).

Indirect effects. Although this could include an almost infi nite number of contribu-tions, here it will be defi ned as including emissions from anything people take with them into the museum (e.g., clothes). Anything more detached, for example, emissions from road transport or fi xed human equipment (e.g., photocopiers) will not be considered.

The effect of people on humidity in the indoor atmosphere is to increase it. People emit water vapor by breathing and the impact of this water release on the humidity will depend on the humidity buffering capacity and air exchange rate of the gallery, as well as the function of any air conditioning or air handling unit installed. The impact of people on a gallery atmosphere can be modeled straight-forwardly (Ankersmit, Tennent and Watts, 2005 ) given some knowledge of the volumes of the galleries, etc. Some museums use exhaled CO 2 as the trigger for temperature or humidity control systems. Possible impacts of humidity have been discussed in Section 12.2 .


Temperature is also increased by people; heating engineers tend to assume that each person in a building is equivalent to just under half a kilowatt of heating. The more interesting effect of visitor fl ows is the chemical source to the atmosphere that they represent. For some pollutants (particularly reduced sulfur and some VOCs as well as particles and dust), people or their clothes can represent signifi cant in - situ sources and transportation routes (Brimblecombe, Shooter and Kaur, 1992 ; Nazaroff et al. , 1993 ; Brimblecombe and Yoon, 2000 ; Yoon and Brimblecombe, 2001 ; Lloyd et al. , 2002 ; Ankersmit, Tennent and Watts, 2005 ).

12.9 Risk Assessment and Preservation Strategies

Any measurement of a museum ’ s atmospheric component should be done for a reason, as part of an atmospheric management plan, or artifact conservation and management strategies. The process of measuring the concentration or varia-tion of that component will differ in methodology with the species being mea-sured, but the process is the same for all. It starts with the reason for the measurement. In museum institutions, in most situations this is the occurrence of damage, a strange odour or some other type of problem. It proceeds through choice of an appropriate person or organization to do the work and to develop a measuring plan and measuring methods to be applied. Having the results in hand and, providing the process has been managed appropriately, it is at this point that the potential diffi culties arise. A raw concentration measurement has great potential to mislead – the interpretation is all. For example, if the concentra-tion of species A in a display case is 500 ppt, is this high or low? If for the sake of argument it is low in comparison to the standard or management plan level, is this because ambient levels are indeed low, or is it instead that the concentra-tion in the display case is being kept low by reacting with the artifacts? It is often the case that other information is required to interpret chemical measurements. In the situation above it is the Case Exchange Rate that is the crucial information which will allow interpretation of the raw measurement of the concentration of A. Care at the mid - point of the process allows intelligent and cost effective inter-vention to be planned and implemented. Guidance on measurements and their interpretation in the context of cultural heritage is shortly forthcoming from CEN/TC 346, a specialist European group concerned with the measurement of ambient air.

At present, the development of adequate monitoring strategies is one of the focal points. Because of the complex nature of pollutants and infl uencing parameters the link between pollutant concentrations detected and the risk potential on objects is not yet known and fully understood. Therefore, easily applicable directives as basic principles of preventive conservation are in the foreground (avoid - block - detect - respond - recover/treat) (T é treault, 2003 ; Grzywacz, 2006 ).

12.10 Conclusion 293

2) Developed by the ad hoc IRK/AOLG working group, obtained from www.umweltbundesamt.de.

12.9.1 Recommendations and Guidelines

An important stage in the growth of this area is the identifi cation of frameworks, defi nitions and the forms of guidelines or standards for indoor air quality from the perspective of cultural heritage. The diffi culty of developing an adequate scheme of evaluation based on the dosage of pollutants received by the artifacts and their conservation management strategy is now beginning to be addressed (T é treault, 2003 ; Ankersmit, Tennent and Watts, 2005 ).

In contrast to the fi eld of indoor air hygiene and indoor air quality (health) where several guidelines have been established over recent years, (e.g., in the UK COMEAP, 2004 and Guideline value recommendations in Germany 2) ) recommen-dations or guidelines for indoor air quality in terms of cultural heritage do not yet exist. The indoor environment in specifi c museum institutions and the parameters affecting museum exhibits are too manifold and the sensitivity of artifacts – depend-ing on the object materials and on previous preservation and restoration treat-ments, respectively – is too different to establish general recommendations. However, in the last few years T é treault (2003) has tried to develop an evaluation scheme by transferring a toxicological concept to the conservation fi eld to estimate exposure - effect relationships for artifacts. For a number of compounds the author has published NOAEL, LOAEL and LOAED - values. Table 12.4 gives an overview of the common museum pollutants with notes on sources, typical levels and stan-dard values for them if applicable. The evaluation of detected values only on the basis of this concept is diffi cult as it does not bear the complex parameters in the museum environment in mind, in particular synergistic effects. A debate on prin-ciples of this evaluation scheme has not been carried out until now.

12.10 Conclusion

Different pathways and sources contribute to a variety of inorganic and organic compounds in the museum environment, whose special position among indoor rooms is founded by the artifacts which can either act as emissions sources them-selves or get damaged by interactions with airborne pollutants. Due to the com-plexity of chemical substances and climatic parameters infl uencing chemical reactions among both pollutants and indoor air and artifacts, not all conceivable damages on artifacts and secondary reactions are known. These processes as well as establishing risk assessment strategies and developing recommendations (e.g., for the use of construction materials, handling of pollution induced damages) will be major tasks for future research.

294

12 Indoor Pollutants in the Museum

Environment

Table 12.4 Common museum pollutants with notes on sources, typical levels and standard values for them if applicable.

Pollutant Source code Notes Typical levels Standard or guide values a,b

Carbonyl sulphide (OCS)

1,2,3,6 Ankersmit, Tennent and Watts (2005) Graedel et al. (1985) T é treault (2003) Brimblecombe, Shooter and Kaur (1992)

Depends on location 400 – 900 ppt. Clean ∼ 550 ppt.

LOAED ∼ 5 μ g m − 3 a − 1 For of CP 5 years LOAED ⇒ [OCS] 1 ⇒ 660 ppt NB with OCS+H 2 S

Hydrogen sulphide (H 2 S)

5,7 Ankersmit, Tennent and Watts (2005) TRL and UK Highways Authority (Personal Communication) T é treault (2003) Brimblecombe, Shooter and Kaur (1992)

< 80 ppt clean; typical galleries 150 ppt; Victoria, London (street) 7000 ppt

LOAED ∼ 1 μ g m − 3 a − 1 For of CP 5 years LOAED ⇒ [H 2 S] 1 ⇒ 130 ppt NB with OCS+H 2 S

Ozone (O 3 ) 3,4,7 http://www.ace.mmu.ac.uk/eae/english.html Felix et al. (2006) Jakobi and Fabian (1997) T é treault (2003)

Outdoor Clean 15 – 40 ppb polluted Europe 60 – 90 ppb (USA 200 ppb) ppb. Indoor 40 – 120 ppb Sao Paulo. I/O ratios 0.3 – 0.5 Annual average UK (20 ppb).

LOAED ∼ 0.005 – 400 μ g m − 3 a − 1 . See T é treault (2003) . For of CP 5 years LOAED ⇒ [O 3 ] 0.005 ⇒ 0.5 ppt ! 1 ⇒ 90 ppt 100 ⇒ 9 ppb

Nitrogen dioxide (NO 2 )

2,3 http://www.airquality.co.uk/archive/ http://www.epa.gov/oar/aqtrnd99/ Garcia Algar et al. (2004) Hisham and Grosjean (1991)

Annual average 0.5 – 3 ppb UK remote,15 – 60 ppb urban. US cities < 80 ppb. 6 – 24 ppb inside homes UK and Spain. 20 – 65 ppb outside and 10 – 45 ppb inside USA museums.

LOAED ∼ 1 – 100 μ g m − 3 a − 1 . See T é treault (2003) For of CP 5 years LOAED ⇒ [NO 2 ] 1 ⇒ 0.2 ppb! 100 ⇒ 10 ppb.

12.10 C

onclusion 295

Pollutant Source code Notes Typical levels Standard or guide values a,b

Nitric acid (HNO 3 )

4,8 Hisham and Grosjean (1991) Fischer et al. (2003) Salmon et al. (1990)

3.5 – 13 ppb US outside. 1 – 8 ppb US inside. < 1 ppb inside 0.05 – 0.75 ppb inside.

LOAED ∼ 0.1 – 10 μ g m − 3 a − 1 . See T é treault (2003) . For of CP 5 years LOAED ⇒ [HNO 3 ] 1 ⇒ 7 ppt ! 10 ⇒ 700 ppt.

Sulfur dioxide (SO 2 )

2,3 http://www.airquality.co.uk/archive/ Wilson (1968) http://www.hc - sc.gc.ca/ewh - semt/pubs/air/exposure - exposition/non - carcino_e.html The air quality strategy for England, Scotland, Wales and Northern Ireland – working together for clean air. DETR (2000)

Annual average 0.1 – 1 ppb UK remote, 2 – 4 ppb urban. I/O 0.2 – 1.2 ppb UK homes. USA urban 5 ppb

LOAED ∼ 10 – 1000 μ g m − 3 a − 1 . See T é treault (2003) . For of CP 5 years LOAED ⇒ [SO 2 ] 10 ⇒ 0.7 ppb 1000 ⇒ 70 ppb UK Air Quality Strategy 24 hour mean limit 125 μ g m − 3 (80 ppb) met in UK. WHO air quality guideline 0 ppb. 0 ppb. USEPA standard 0 ppb. Canadian Exposure guidelines for residential air quality ALTER 25 ppb.

Organic acids (Acetic acid, formic acid)

3,4,8 Watts (2007) private communication Grzywacz and Tennent (1994) Gibson (2005) private communication T é treault (2003) Gibson et al. (1997a, b)

Remote < 2 ppb; urban 2 – 7 ppb; in museums galleries < 100 ppb cases < 350 ppb

LOAED ∼ 10 – 5000 μ g m − 3 a − 1 . See T é treault (2003) . For of CP 5 years LOAED ⇒ [OA] 10 ⇒ 0.5 ppb 1000 ⇒ 300 ppb

a T é treault (2003) has produced an excellent reference work on recommended levels or standards. Where indicated check this for specifi c artifact types. b In all cases depends on the artifact, the artifact management plan and conservation period (CP).

Source Codes: (1) permanent atmospheric component; (2) formed from combustion or high temperature; (3) externally sourced; (4) internally sourced; (5) people, bacteria or animals (6) clothes (7) electrical discharge (8) building or internal materials (9).


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Indoor Organic Chemistry Glenn Morrison

13

13.1 Introduction

The building envelope encompasses a unique micro - environment, complete with its own distinctive chemistry. Only recently have reactions among organic and inorganic compounds in the indoor environment been identifi ed and further recognized as important in controlling human exposure to air pollutants. Almost any type of organic reaction found in an organic chemistry textbook occurs indoors, including oxidation – reduction in the ozone – terpene system, hydrolysis of plasti-cizers, or acid – base chemistry at surfaces. Homogeneous chemistry occurs between gas - phase molecules, within solids or even in the thin layers of water that exist on surfaces. Heterogeneous chemistry occurs at the interface of air and fur-nishing surfaces or aerosols. This chapter is organized by the phase in which the reaction occurs and owes much to previous reviews of indoor chemistry (Nazaroff and Weschler, 2004 ; Uhde and Salthammer, 2007 ; Weschler, 2004 ; Weschler and Shields, 1997b ). An outline of the subjects treated in this chapter is shown in Figure 13.1 .

This chapter focuses on the impact of transformative chemistry, rather than more subtle chemical interactions such as sorption. That said, some transforma-tive chemistry will indirectly affect phenomena such as sorption, and is included if it is believed to have a signifi cant impact on the indoor environment. A trans-formation is included if it increases or decreases the gas - phase concentration of an organic compound by an amount likely to produce a signifi cant effect on health or welfare. This includes changes that are a substantial fraction of a promulgated standard. Also included are changes that can infl uence irritation or odor percep-tion (Wolkoff et al ., 2006 ), or changes that are recognized to be associated with an adverse health outcome, such as asthma or cancer.


301

302 13 Indoor Organic Chemistry

13.2 Relevance of Chemistry Using Indoor Air Models

Indoor air quality models that include chemistry and all other important source and loss mechanisms have been developed by a variety of authors (Nazaroff and Cass, 1986 ; Weschler, Shields and Naik, 1989 ). These models can be used to determine if a relevant change is likely to occur, by balancing pollutant source rates with removal rates to arrive at a predicted indoor concentration. The key components are shown here for a steady - state, ‘ well - mixed ’ mass - conservation model of a building:

CE V E V C

kp s o=

+ ++

λλ

where C is the concentration of the compound of interest, E p is the primary emis-sion rate of the compound into the building from all sources, E s is the secondary emission rate of the compound into the building as a result of transformative chemistry, V is the volume of the building, λ is the building air exchange rate (ACH), C o is the outdoor concentration of the compound and k is the sum of all fi rst - order removal rate constants for that compound. Although somewhat simpli-fi ed, this model captures most of the relevant processes and allows for comparison of rates and their infl uence on indoor concentrations.

Figure 13.1 Outline of organic chemistry known to occur indoors.

13.3 Homogeneous Chemistry 303

If chemistry results in a fi rst - order removal rate constant ( k ) that is a substantial fraction of the air exchange rate ( λ ), then this chemistry may be an important removal mechanism. With air exchange rates ranging from 0.1 to 2 h − 1 , the sum of all fi rst - order rate constants ( k ) needs to be greater than ∼ 0.03 h − 1 to signifi cantly decrease indoor reactant concentrations. As will be demonstrated in the main body of the chapter, homogeneous reactions between ozone and many fragrances are fast enough to result in a substantial decrease in ozone and the fragrance. Like-wise, heterogeneous reactions of ozone with oils on the surface of new carpet and other furnishings result in large fi rst - order removal rates of ozone.

Similarly, if the emission rate of the compound due to chemistry, E s , is a sig-nifi cant fraction of primary emission rates or the rate of entry by ventilation, then it is worth considering. Increases in reaction products may be relevant even if rate constants are small. The pseudo fi rst - order rate constant for the reaction of · OH with gas - phase organics is predicted to be small (Nazaroff and Weschler, 2004 ), but the reaction has been shown to be responsible for much aerosol generation (Fan et al ., 2003 ).

13.3 Homogeneous Chemistry

Chemistry that takes place in a phase of uniform composition is termed homoge-neous chemistry. In indoor environments, this chemistry takes place in air, water, coatings, and adhesives.

13.3.1 Gas - Phase Organic Oxidation Chemistry: Ozone

It is fi tting to begin the discussion with one of the most well - studied reactions in indoor air quality science. In 1999, Weschler and Shields (1999) showed that ozone will react with some terpenes at substantial rates in indoor environments. This prompted phenomenological, health, theoretical, kinetic and mechanistic studies to understand the potential and real impact of this chemistry on occupants.

Terpenes form a large and varied class of biogenic VOCs that all derive from the base unit, isoprene. Terpenes are responsible for the fresh scent of lemons, pine, and camphor. From the isoprene building block, hundreds of isomeric forms of monoterpenes (C 10 H 16 ), sequiterpenes (C 15 H 24 ) and others can be formed in plant tissues. They are found naturally in plant - based materials but are commonly added to consumer products as fragrance. For example, monoterpenes are emitted into buildings from many sources including wood products (Hodgson, Beal and McIlvaine, 2002 ), cleaners (Nazaroff and Weschler, 2004 ) and air fresheners (Sarwar et al ., 2004 ). Limonene, α and β - pinene, 3 - carene, isoprene and p - cymene are among the more commonly detected terpenes in surveys of VOCs in indoor air (Brown et al ., 1994 ; Hodgson and Levin, 2003 ).


Many terpenes contain one or more double bonds that are readily attacked by ozone. The relevance of this class of reactions can be established by considering the kinetics of the α - pinene – ozone reaction. As the name implies, α - pinene is a fragrance that smells of pine, is emitted from wood building products (Hodgson and Levin, 2003 ), cleaners (Nazaroff and Weschler, 2004 ) and air fresheners (Salthammer, 1999 ; Sarwar et al ., 2004 ; Singer et al ., 2006b ). Select unsaturated indoor compounds, their sources and their pseudo - fi rst - order rate constants are shown in Table 13.1 , and a more extensive listing is in Nazaroff and Weschler (2004) . α - pinene reacts with ozone with a second - order rate constant equal to 2.1 × 10 − 6 ppb − 1 s − 1 . Given a typical indoor mixing ratio of ozone of ∼ 20 ppb (Weschler and Shields, 2000 ), a pseudo - fi rst - order rate constant for this system is 0.15 h − 1 as shown in Table 13.1 . The magnitude of this rate constant is comparable to the low end of indoor air exchange rates and is therefore in the relevant range of interest. Under these conditions, a substantial fraction of α - pinene will be converted to products. Many terpenes known or expected to be present in indoor air are expected to react at relevant rates (Nazaroff and Weschler, 2004 ). Yet under the same conditions, camphene, a terpene commonly added to cleaners and air fresheners, is slow to react. Concerns related to atmospheric chemistry have resulted in most of the experimental determination of these kinetic rate constants (Atkinson and Arey, 2003 ). The determinations of the rate constants for α - terpineol (Wells, 2005 ), citronellol (Ham, Proper and Wells, 2006 ), citronellal (Harrison, Ham and Wells, 2007 ), geraniol (Forester, Ham and Wells, 2007 ), and dihydro-myrcenol (Forester, Ham and Wells, 2006 ) are unique in that these are early instances of detailed kinetic analysis driven entirely by indoor concerns. Species shown in Table 13.1 are commonly odorous at the concentrations observed in indoor environments, as they are usually used as fragrances. Thus, by defi nition, a reduction in fragrance intensity is ‘ relevant ’ by our standards. Similarly, the

Table 13.1 Select terpenes and their pseudo - fi rst - order rate constants (h − 1 ) with O 3 , · OH , and NO 3 . a

Compound k ′ , O 3 @ 20 ppb

k ′ , · OH @ 10 − 5 ppb

k ′ , NO 3 @ 10 − 3 ppb

Sources

Camphene 0.002 0.048 0.016 cleaners and air fresheners Styrene 0.03 0.070 0.004 carpet, adhesives α - pinene 0.15 0.048 0.55 cleaners and air fresheners Limonene 0.37 0.15 1.1 cleaners, air fresheners and perfumes Citronellol b 0.43 0.15 NA cleaners, insect repellants α - terpineol c 0.54 0.17 NA cleaners, pine oil, mold Linalool 0.76 0.14 1.0 air fresheners, perfumes

a Except where noted, values adapted from table 8 of Nazaroff and Weschler (2004) . b Value304s adapted from Ham, Proper and Wells (2006) . c Values adapted from Wells (2005) .


Scheme 13.1 Selected products and yields ( ) of the reaction of α - pinene with ozone.

occupant benefi ts from this chemistry when styrene, a carcinogen, and ozone, a potent respiratory irritant, are both reduced in concentration.

While a comparison of rates can demonstrate relevance, indoor ozone chemistry becomes even more interesting when pathways and products of the oxidation are studied in detail, such as for the α - pinene – ozone reaction (Scheme 13.1 ). Ozone adds across the ring - bound double bond forming a primary ozonide. This


intermediate rearranges, or fragments, into many products; only a few are shown in Scheme 13.1 to demonstrate the range and consequences. Formaldehyde, a carcinogen and irritant, is formed with a ‘ yield ’ of ∼ 0.15 (Atkinson and Arey, 2003 ) (and references therein). This means that for every 100 molecules of ozone that reacts with α - pinene, about 15 molecules of formaldehyde will form. Yields depend on the structure of the terpene; β - pinene, has a formaldehyde yield of ∼ 0.7. Form-aldehyde yields for terpenes range as high as 0.9 (isoprene). Destaillats et al . (Destaillats et al ., 2006b ) showed that, when relatively high levels of ozone react with terpenes emitted from cleaners, formaldehyde rises to levels (13 – 35 ppb) much greater than the California reference exposure level of 2 ppb.

Carbonyl compounds are irritants and heavier carbonyls tend to have lower irritation thresholds (Cometto - Mu ñ iz, Cain and Abraham, 1998 ). Suspect irritants include pinonaldehyde, a stable di - aldehyde that is formed at high yield in the ozone – α - pinene reaction. Anderson et al . (2007) found, through the application of quantitative structure activity relationships and animal models, that most dicar-bonyl compounds are irritants and sensitizers. Organic acids tend to be roughly 10 times as irritating as their analogous aldehydes; a number of acids (formic acid), di - acids (pinic acid) and acid/aldehyde compounds (norpinonic acid) are also formed. Intermediate products, such as ozonides, and hydroperoxy radicals are highly reactive and may also interact strongly with mucous membranes.

Suspicions that ozone – terpene reaction products might be irritating or worse, prompted several human and animal studies. Early mouse studies of reactions of ozone with isoprene (Wolkoff et al ., 2000 ), α - pinene (Wolkoff et al ., 1999 ) and D - limonene (Clausen et al ., 2001 ), identifi ed products that acted as airway irritants. Neither the individual reactants nor the aged products were as irritating as the mixture during reaction. Further, the more aged reaction mixtures were less irri-tating than younger mixtures (Wilkins et al ., 2003 ), suggesting that unidentifi ed products, such as short - lived radicals, were responsible for the irritation. Respira-tory fl ow was limited in mice exposed to reaction products, and possible long - term sensitization was observed (Rohr et al ., 2002 ). As evidence of eye irritation, the limonene – ozone reaction mixture caused eye - blink frequency to increase in human subjects (Kleno and Wolkoff, 2004 ). Tam á s et al . (2006) recently evaluated perceived air quality using 20 human subjects exposed to a limonene – ozone reac-tion mixture in an offi ce environment. They found that the sensory load for the reaction mixture was much greater than for either compound alone. In contrast to these studies, a recent study incorporating 130 female human subjects found no signifi cant subjective or objective health effects of a mixture of ozone and VOCs (including α - pinene and D - limonene) (Fiedler et al ., 2005 ). Even though aldehydes, peroxides, acids and aerosols were formed by these reactions, only stress was shown to be correlated with adverse health effects. Similarly, Laumbach et al . (2005) did not observe signifi cant increases in markers of infl ammation in nasal lavage fl uid from the same 130 subjects.

Oxidation of terpenes creates low - vapor pressure species that self - nucleate to form small aerosols, or condense on and increase the mass of existing aerosols. These secondary organic aerosol s ( SOA s) are of concern because they can signifi -


cantly increase indoor concentrations of respirable particulate matter. Weschler and Shields (1999) studied particle generation due to terpene – ozone reactions in two adjacent, identical offi ces. Using one offi ce as a control, they found statistically signifi cant increases in submicron particle concentrations in the offi ce with ele-vated ozone and either limonene, α - terpinene or a terpene mixture from a cleaner. For experiments using outdoor ozone (instead of injected ozone), particle concen-trations followed the rise and fall of indoor ozone concentrations. Long, Suh and Koutrakis (2000) quantifi ed particle generation in fi eld homes and showed that a pine - scented cleaner increased submicron particle concentrations in the presence of infi ltrated ozone. In a laboratory chamber, Wainman et al . (2000) studied the growth of submicron particles due to the limonene – ozone reaction. They observed a 7 - to 100 - fold increase in the number concentration of particles with diameters less than 0.5 μ m. (Rohr et al ., 2003 ). Sarwar et al . (2004) showed, in laboratory chamber experiments, that SOAs increase substantially when realistic levels of ozone combine with terpenes from cleaners, air fresheners and perfumes. Shown in Figure 13.2 is an experiment (Sarwar et al ., 2004 ) in which a solid air freshener is introduced into a chamber with an air exchange rate of 0.62 h − 1 , and with ozone initially at ∼ 110 ppb. A sharp increase in aerosols occurs due to the reaction of the terpene mixture and ozone, and is sustained for several hours. Note that the reported particle mass concentration rises above outdoor standards. Hubbard et al . (2005) and Waring, Siegel and Corsi (2008) demonstrate that ozone generat-ing particle fi lters, advertised as ‘ air cleaners ’ designed to reduce indoor particle concentrations, actually increase particle mass substantially in the presence of terpene fragrances.

The magnitude of aerosol generation is suffi cient to make these reactions of interest in regard to human health. A large body of epidemiological literature indicates that increases in ambient aerosol concentrations are associated with increased mortality. In particular, an increase of 25 μ g m − 3 in particle mass concentration for PM2.5 (total particle mass for particles with an aerodynamic

Figure 13.2 Aerosol formation due to ozone reactions with terpene emitted from a solid air - freshener in a large laboratory chamber. Adapted from (Sarwar et al ., 2004 ).


diameter < 2.5 μ m) is correlated with a 1 – 3.5% increase over baseline mortality and a 1 – 12% increase in hospitalization for respiratory and cardiovascular disease (USEPA, 2004 ). For typical indoor ozone concentrations, in the presence of a pine cleaner, Weschler and Shields (1999) observed a 15 – 20 μ g m − 3 increase in particles < 1.0 mm in diameter. Similarly, Wainman et al . (Wainman et al ., 2000 ) found that a 20 μ g m − 3 increase due to these reactions is probable under typical indoor condi-tions. Where ozone is introduced by an ozone generating air fi lter in fi eld homes, Hubbard et al . (2005) observed a 10 – 20 μ g m − 3 increase in particles < 1.0 mm in diameter due to the ozone reaction with a variety of terpene sources. The more extreme, but not unrealistic, conditions simulated by Sarwar et al . (2004) in a chamber show that these reactions can result in particle mass concentrations (for 0.02 to 0.7 μ m diameter particles) greater than ambient PM2.5 standards (Figure 13.2 ). Increasing ventilation rates tends to decrease SOA concentrations by dilu-tion, but also tends to shift the respirable particle size distribution to smaller sizes (Weschler and Shields, 2003 ) because the particles have less time to grow or coagulate.

Although these reactions result in an increased respirable aerosol mass, little is known about human health effects associated with these particular condensed reaction products. Tam á s et al . (2006) observed a correlation between the number concentration of SOAs, from the limonene – ozone reaction in an offi ce, and sensory load reported by 20 human subjects. Additionally, a screening study indi-cates that perfume wearers will be subjected to a ‘ personal reactive cloud ’ of reac-tion products, such as fi ne aerosols (Corsi et al ., 2007 ).

13.3.2 Gas - Phase Organic Oxidation Chemistry: Hydroxyl Radical

The hydroxyl radical, · OH, formed in the terpene – ozone chemistry discussed thus far, is a much stronger oxidant than ozone and reacts indiscriminantly with almost any hydrocarbon. Hydroxyl radical yields for the ozone – terpene reaction can be quite high, nearly unity for α - pinene (Atkinson and Arey, 2003 ). Based on the prior observations of · OH formation (Atkinson et al ., 1992 ), Weschler and Shields (1996) predicted that · OH could reach ∼ 10 − 5 ppb, greater than outdoor nighttime mixing ratios. They confi rmed this prediction in an offi ce environment by measur-ing · OH formed from the reaction of ozone and limonene, using cyclohexane as an · OH scavenger (Weschler and Shields, 1997a ). By modifying an outdoor atmo-spheric chemistry model, Sarwar et al . (2002) was able to simulate indoor · OH for typical indoor VOC mixtures. They predicted · OH mixing ratios similar to those measured by Weschler and Shields. Singer et al . (2006a) recently showed that the mixing ratio of · OH ranged from 1 − 9 × 10 − 5 for the reaction of ozone with terpenes emitted from a variety of cleaning products. Reactions of ozone with unsaturated organic compounds are thought to be the primary source of · OH in indoor envi-ronments; thus, its further reactions with terpenes are of great interest. That is, if · OH is present, terpenes and similar compounds are probably also present at high concentrations.


Revisiting Table 13.1 , we see that pseudo - fi rst - order rate constants for the reac-tion of · OH (at 10 − 5 ppb) with unsaturated organic compounds are much lower than that for ozone. This is because, even though · OH yields are high, the hydroxyl radical concentration is much lower than ozone (Sarwar et al ., 2002 ; Weschler and Shields, 1996 ; Weschler and Shields, 1997b ). These removal rates by reaction with · OH are relevant primarily for buildings with low air exchange rates ( ∼ 0.1 h − 1 ) (Nazaroff and Weschler, 2004 ). Yet, important products are formed. In the pres-ence of nitric oxide, the hydroxyl radical can generate organic nitrates with yields as high as 0.18 for α - pinene (Noziere, Barnes and Becker, 1999 ). Organic nitrates, specifi cally peroxyacetylnitrate, are potent irritants. Fan et al . (2003) showed that · OH, generated from ozone – terpene reactions in the presence of 21 saturated VOCs, was responsible for most of the formaldehyde and p - tolualdehyde produc-tion and 19 – 29% of the organic aerosol formation. The hydroxyl radical can also enhance concentrations of products initially formed in the ozone reaction. In the reaction of ozone with D - limonene, Sarwar et al . (2002) predicted that · OH is responsible for about a 4 - fold increase in 3 - isopropenyl - 6 - oxoheptanal (over ozone alone). They also predicted roughly a 2 - fold increase in pinaldehyde due the reac-tion of · OH with α - pinene, over ozone alone.

13.3.3 Gas - Phase Organic Oxidation Chemistry: Nitrate Radical

The nitrate radical, NO 3 , has also been identifi ed as a potent agent of change in indoor environments. The nitrate radical forms as the product of the reaction of O 3 with NO 2 and will decompose rapidly in daylight. However indoor lighting is insuffi ciently intense to photolyze NO 3 ; this allows indoor concentrations to reach part - per - trillion levels. The nitrate radical readily adds to double bonds, making terpenes easy targets for nitration (Kirchner and Stockwell, 1996 ). As Weschler and Shields (1997b) noted, the combination of low lighting and high concentra-tions of unsaturated VOCs makes this chemistry potentially important in indoor environments. Indeed, organic nitrates have been detected at substantial yield in the reactions with isoprene and α - pinene (Atkinson and Arey, 2003 ). With a focus on the important indoor terpenoid α - terpineol, Jones and Ham (2008) observed the formation of several organic nitrate species, some of which were determined to be ‘ borderline ’ or ‘ undoubtedly hazardous ’ . Many of the same radical intermedi-ates associated with ozone and · OH chemistry have been observed as the result of NO 3 · terpene chemistry, including the peroxy radical, HOO · , and secondary aerosols. Beyond terpenes, Nazaroff and Weschler (2004) specifi cally identify the reaction of NO 3 with 2 - butoxyethanol as troubling due to the formation of propyl-nitrate, and other unique products.

Referring to Scheme 13.1 , we see that other radical products are formed in the reaction of α - pinene (and other terpenes) with ozone. Not discussed here are secondary or tertiary reactions that may occur as the result of reactions between VOCs and these radical products. The complexity increases many - fold as we consider the many different kinds of unsaturated VOCs and the many different


saturated VOCs that may react with OH, NO 3 and other secondary oxidants. Overall, product yields depend strongly on the environment and presence of co - reactants. Capouet et al . (2004) predict that pinonaldehyde yields from the α - pinene - · OH reaction are six time higher in a polluted atmosphere than in a more pristine setting. Pinonaldehyde yield is also fairly sensitive to relative humidity (Warscheid and Hoffmann, 2001 ). Product yields for each indoor microenviron-ment may then be as unique as a fi ngerprint. Combined, the complex chemistry here is suggestive of turning a relatively benign fragrance compound into a broad spectrum of oxidized compounds of uncertain human toxicity.

13.3.4 Condensed - Phase Chemistry: Oxidation

A great deal of chemical transformation takes place before consumers receive their furniture, fl ooring, and other indoor building materials. Polymerization of glues, coatings and plastics, or heat processing of manufactured wood products, gener-ates volatile by - products that can continue to be emitted from the material after it has been installed. Once installed, further transformations increase the load of odorous or toxic gas - phase species (Uhde and Salthammer, 2007 ). Reactions dis-cussed here include oxidation and hydrolysis.

The auto - oxidation and photo - oxidation of fatty acids and triglycerides is a well - studied phenomenon (rancidity), and has been recognized to occur indoors in linoleum (Jensen, Wolkoff and Wilkins, 1996 ), furniture coatings (Salthammer, Schwarz and Fuhrmann, 1999 ) and in cooking oils. This oxidation results in the formation of volatile aldehydes, methyl esters, alcohols and more. See Scheme 13.2 for oxidation chemistry of linolenic acid, a major constituent in linseed and other oils. The molecule is truncated to show only the last nine carbons of the 18 carbon chain. Oxidation is initiated when hydrogen is abstracted, generating a radical - stabilized resonance structure that may shift the radical position along the carbon backbone of the fatty acid. This is followed by addition of oxygen to the radical location and abstraction of hydrogen from a neighboring ester, propogating the oxidation chemistry from one ester to another. The resulting hydroperoxy compound is semi - stable, but may eventually decompose forming a variety of short - chain, volatile, compounds such as odorous aldehydes and carboxylic acids.

The relevance of this chemistry is demonstrated by the sensory impressions of linseed - oil based linoleum. Jensen, Wolkoff and Wilkins (1995) studied the auto - oxidation products emitted from linoleum and identifi ed saturated and unsatu-rated aldehydes and fatty acids. An odor evaluation of the identifi ed 2 - alkenals and the fatty acids as contributing most to odor intensity (Jensen, Wolkoff and Wilkins, 1995 ), although many odorous products may not be quantifi ed by traditional ana-lytical methods (Knudsen et al ., 2007 ) Knudsen et al . (2007) showed that linseed oil based products exhibited a more negative sensory perception than similar product not containing linseed oil, and that the negative perception persisted for at least a year.


13.3.5 Condensed - Phase Chemistry: Hydrolysis

Excessive dampness in buildings can lead to a host of problems, mold growth being one of the more obvious ones. However, water also promotes hydrolysis of glues, plasticizers and fl ame retardants. A classic example is the hydrolysis of urea – formaldehyde resins with the carcinogen formaldehyde as a product (Scheme 13.3 ). These resins are widely used in manufactured wood products (plywood, particle board and so forth). In western countries, manufacturers have reduced formaldehyde emission rates by reducing the amount of formaldehyde in the resin, adding formaldehyde scavengers and using coatings as barriers to emissions (Conner, 1996 ). However, formaldehyde continues to be a problem in residences in China, perhaps due to emissions from furnishings (Wang, Lee and Ho, 2007 ).

As shown in Scheme 13.3 , hydrolysis has the effect of breaking down condensed materials to more volatile molecules. Uhde and Salthammer (2007) explain that

Scheme 13.2 Auto - oxidation of linolenic acid (truncated) leading to radical propagation, intermolecular cross - linking, and volatile aldehyde emissions.


esters are cracked in this way and converted to alcohols and carboxylic acids. Esters known to hydrolyze indoors include the plasticizers DEHP and di/iso - butylphthal-ate (DBP/DIBP). DEHP forms the odorous and volatile 2 - ethyl - 1 - hexanol by hydro-lysis (Scheme 13.4 ).

Although hydrolysis is generally slow, Uhde and Salthammer (2007) explain that by installing PVC fl ooring on poorly cured concrete, the secondary emission rate of 2 - ethyl - 1 - hexanol is enhanced. The fl ooring traps moisture and the alkaline concrete catalyzes the hydrolysis. Weschler (2004) has suggested that observed associations between damp buildings and health complaints Bornehag et al ., 2001 ) may be due to products of hydrolysis chemistry. In support of this hypothesis, Bornehag et al . (2005) observed that PVC fl ooring, in combination with water leakage in residences, was associated with a higher prevalence of childhood respi-ratory complaints such as wheezing and asthma.

Hydrolysis of fl ame retardants can generate a variety of odorous compounds, including halogenated alcohols. Halogenated and nonhalogenated organophos-phates have been added to building materials such as polyurethane construction foam. Products of organophosphate hydrolysis were detected by Salthammer, Fuhrmann and Uhde (2003) from recycled fl ooring materials, including 1 - chloro - 2 - propanol and 2 - chloro - 1 - proponol. Hydrolysis of tris(1,3 - dichloro - 2 - propyl)phos-phate ( TDCPP ) is shown in Scheme 13.5 (Uhde and Salthammer, 2007 ). The product, 1,3 – dichloro - 2 - propanol is considered to be a human carcinogen by the Deutsche Forschungsgemeinschaft (German Research Foundation) based on animal studies. Newer products contain primarily nonhalogenated organophos-phates, due to concerns over the toxicity of halogenated fl ame retardants. However, recycled materials retain the older, chlorinated compounds (Salthammer, Fuhrmann and Uhde, 2003 ).

Scheme 13.3 Hydrolysis of urea – formaldehyde resin, generating formaldehyde.

Scheme 13.4 Hydrolysis of diethylhexylphthalate to form 2 - ethyl - 1 - hexanol.

13.4 Heterogeneous Chemistry 313

13.4 Heterogeneous Chemistry

Chemistry that occurs at interfaces is termed heterogeneous chemistry and is remarkably important despite the tiny volume taken up by the thin layer of mol-ecules coating indoor surfaces (Morrison, 2008 ). This is due to the following: (i) the available surface area is large relative to the total volume of a typical building; (ii) surface sorption extends the average residence time of reactants and increases the probability that conversions will occur; (iii) unique compositions and mor-phologies at surfaces can promote some reactions (that is, catalytically) or promote selectivity in reaction pathways. Condensed phases are roughly 1000 times more concentrated (by mass) than the overlying gas, meaning that reactions taking place at the surface can proceed with speed. Indeed, it is common for surface reactions to be nearly instantaneous but limited by the availability of a gas - phase reactant such as ozone.

13.4.1 Heterogeneous Chemistry: Ozone and Fresh Indoor Surfaces

Ozone is once again an important reactant here, as it tends to react rapidly at indoor surfaces. The primary loss mechanism for ozone in indoor volumes is surface reactions, causing the I/O ratio of ozone to be typically 0.2 or lower (Weschler, Shields and Naik, 1989 ). Using a mass - conservation approach, and assuming that the outdoor mixing ratio is at the US 24 - h standard (80 ppb) and that every reacted ozone molecule creates a volatile product, then the total indoor mixing ratio of products could be of the order of ∼ 0.8 × 80 ppb or 64 ppb. This is substantial for the types of products that tend to be generated: aldehydes, carbox-ylic acids, nicotine oxidation products, and, indirectly, aerosols.

Ozone chemistry has been studied on nearly every relevant surface in indoor environments. Ozone uptake rates on surfaces, regardless of the chemistry, have been quantifi ed on carpets, tiles, concrete, wood, and glass. Because this chapter targets organic chemistry, the focus will be on those surface reactions that involve organic compounds. Ozone can react rapidly on the surface of a variety of virgin materials due to presence of unsaturated organic species remaining in or on

Scheme 13.5 Flame retardant TDCPP hydrolysis forming 1,3 - dichloro - 2 - propanol.


surfaces after production, including carpet, paint, and materials used in heating ventilating air conditioning system s ( HVAC s).

Carpet In a seminal study of indoor chemistry, Weschler, Hodgson and Wooley (1992) found that ozone reacted with carpet surfaces generating formaldehyde, acetaldehyde and C4 − C10 aldehydes. A key fi nding was that ozone converted some of the condensed phase material on the carpet to volatile species, resulting in a large increase in gas - phase VOC mass. Further investigations (Morrison and Nazaroff, 2002 ) found that ozone prompted the formation of C 1 − C 13 n - aldehydes, as well as unsaturated aldehydes, on new carpet. Based on product yield studies, 2 - nonenal was predicted to signifi cantly exceed odor thresholds under typical conditions and that odorous conditions could persist for years. Field measure-ments of secondary emissions from carpet (Wang and Morrison, 2006 ) showed that one - to two - year - old carpet generated substantially more secondary aldehydes than carpet that had been installed for over 10 years. Thus, secondary emissions can persist for at least two years, but surface reactants may eventually become depleted. Knudsen et al . (2003) performed sensory evaluations on carpet and other materials that had been exposed to ozone. They found that carpet in particular exhibited signifi cantly higher odor intensity as a result of secondary emissions.

Latex Paint Paint and other interior surfaces can also generate secondary car-bonyl species. Reiss et al . (1995a) studied ozone reactions on latex paint and found that many of the paints will generate formaldehyde as a secondary species. Ozone also induced acetone and acetaldehyde emissions from a smaller subset of the paints, but well - aged paint was not as reactive. They estimated that ozone reactions on newly applied paint would constitute 10 – 15% of the total emission rate of formaldehyde measured in fi eld homes (Reiss et al ., 1995b ). By isolating wall sec-tions in fi eld homes and exposing the sections to ozone, Wang and Morrison (2006) observed secondary emissions of aldehydes from latex painted walls, which were low relative to other surfaces tested.

HVAC Materials Ventilation duct liners also react with ozone forming formalde-hyde, acetone and C 5 − C 10 aldehydes. Morrison et al . (1998) subjected new and used duct liners, air fi lters, sealants, sheet metal and other HVAC materials to ozone in small chambers. They observed secondary emissions of C 5 − C 10 aldehydes from a new duct liner, a neoprene gasket and duct sealants. They predicted that secondary emissions from these materials could increase indoor aldehyde concen-trations to levels comparable with odor thresholds. As will be discussed later, soiled HVAC materials also generate secondary products.

On indoor surfaces, aldehydes are the prominent product observed, probably because they are more volatile and easier to detect than other possible products. In Scheme 13.6 , ozone splits the linoleic ester into C 6 and C 9 aldehydes and acids. In ozone reactions with carpet, Morrison and Nazaroff (2002) found the aldehyde yield to range from 0.1 to 0.74 for summed aldehydes. A yield of 0.5 would be expected if aldehydes and acids were formed equally in Scheme 13.6 . Low yields


may indicate that surface conditions favor the formation of acids or other low - volatility products, or that other reactions consume ozone (such as homogenous decomposition). High yields suggest that surface conditions favor aldehydes. They also detected alkanes that have been reported to form as products of the decom-position of the ozonide (Gunstone, 2004 ). Interestingly, lower - volatility products may also contribute to aerosol growth by partitioning from the indoor surface to existing aerosols (Aoki and Tanabe, 2007 ; Beko, Clausen and Weschler, 2007 )

Does the composition of an indoor surface infl uence the product yields for this chemistry? To date, no clear evidence of this behavior has been reported, but industrial and aerosol chemistry reports are suggestive. Ozonolysis (the treating with high concentrations of ozone) of fatty acids/esters has been used industrially to produce aldehydes, acids, esters, alcohols and low - volatility resins. The products of ozonolysis depend on the location of double bonds in the fatty acid and the presence of solvents, catalysts, co - reactants and moisture (Gunstone, 2004 ). For example, the presence of peroxyacids encourages acid formation. Ammonia in the presence of a zinc catalyst encourages amine formation. Recent investigations of aerosol chemistry suggest that acids can catalyze the reaction of aldehydes to less volatile products (Jang et al ., 2003 ; Jang et al ., 2002 ). Therefore, soiled indoor surfaces may provide a complex mixture of catalysts, co - reactants, salts and sol-vents that favor (or disfavor) a wide variety of volatile species.

Over time, original surface reactants will eventually become depleted and sec-ondary emissions of by - products should be reduced. Many studies have shown that ozone uptake on indoor surfaces tends to decrease with continued exposure, a phenomenon known as ‘ aging ’ (Morrison and Nazaroff, 2000 ; Morrison et al ., 1998 ; Reiss et al ., 1995a ; Sabersky, Sinema and Shair, 1973 ; Simmons and Colbeck, 1990 ). Further, there is evidence that secondary emission rates also decrease with time. Morrison and Nazaroff (2002) showed that secondary aldehydes on new carpet fi bers, in a fi xed - bed reactor, could be depleted in a day; however, the reac-tivity of whole carpet was not substantially decreased over the relatively short time periods studied. Wang and Morrison (2006) showed that carpet in older homes

Scheme 13.6 Heterogeneous ozone oxidation of linoleic ester to form volatile aldehydes and less volatile acids.


exhibited signifi cantly lower secondary emission rates of aldehydes, than carpet in newer homes.

13.4.2 Heterogeneous Chemistry: Ozone and Soiled Surfaces

Ozone will also react with species that have accumulated on surfaces over time. Cooking oils, soaps, tobacco products and even human skin oils have all been shown to react with ozone to form volatile products. This kind of chemistry is especially interesting, as reactants continue to be supplied by occupant activities; a semi - steady level of chemical reaction and product formation is anticipated (rather than eventual depletion).

Morrison et al . (1998) observed that used duct liners from commercial HVAC systems consumed more signifi cantly ozone than new duct liners. This suggests that dust build - up and soiling deposit reactive compounds over time on the inner surfaces of HVAC systems. In related work, several studies have recently shown that soiled HVAC fi lters are more reactive than clean fi lters (Hyttinen, Pasanen and Kalliokoski, 2006 ; Hyttinen et al ., 2003 ; Zhao, Siegel and Corsi, 2007 ). Hyt-tinen et al . (2003) observed formaldehyde as the only reaction product from a subset of these fi lters. Bek ö et al . (2006) observed that used fi lters that were treated with ozone were perceived as ‘ less acceptable ’ than fi lters treated with plain air or nitrogen, suggesting that objectionable by - products were formed. They suggest that this chemistry may account for previous observations that sick building syn-drome symptoms and poor occupant performance correlate with the presence of loaded HVAC fi lters (Beko, Clausen and Weschler, 2008 ; Wargocki, Wyon and Fanger, 2003 ; Wyon et al ., 2000 ). Fine aerosols have also been observed as indica-tors that semi - volatile reaction products form and partition from the soiled fi lter surface to aerosols (Beko, Clausen and Weschler, 2007 ).

Wang, Springs and Morrison (2005) showed that countertops that become coated with cooking oils or soaps will generate volatile aldehydes in the presence of ozone, in chemistry akin to Scheme 13.6 . They verifi ed in fi eld studies (Wang and Morrison, 2006 ) that kitchen counters are a major emitter of secondary alde-hydes (on a unit area basis).

Wisthaler et al . (2005) and Weschler et al . (2007) studied ozone – surface reactions in a simulated aircraft environment that included used carpet, seats and other inner airplane surfaces. They also evaluated the presence of T - shirts worn by volunteers for a day. A key fi nding was that ozone consumption increases due to reactions with human skin oils, specifi cally squalene, on T - shirts and other air-plane surfaces. Using a very sensitive proton - transfer mass spectrometry system, they observed volatile oxidation products of the ozone – squalene reaction including acetone and 6 - methyl - 5 - heptenone. They identifi ed aldehydes, similar to those observed from earlier carpet studies, which may be generated from reactions with airline carpet and upholstery. They also identifi ed organic acids and organic nitrates as volatile reaction products. Studying discrete pieces of aircraft and cloth-ing materials, Coleman et al . (2008) observed these compounds generated in comparable yields.


Ozone – nicotine surface chemistry has recently been identifi ed by Destaillats et al . (2006a) . Nicotine and other products of tobacco smoking adsorb strongly to indoor surfaces. In the presence of ozone, surface nicotine can form a variety of products (Scheme 13.7 ). On cotton, high humidity inhibits this reaction. Ozone attacks the pyrrolidinic N, suggesting that other indoor surface bound species with this functionality may be oxidized in this manner. Not only does this chemistry generate volatile by - products, it may explain why indoor nicotine concentrations correlate poorly with exposure to tobacco smoke. This raises the question, what other, similarly functionalized, surface amines may also react with ozone in indoor environments?

Homogeneous reactions with terpenes have already been discussed, but surface chemistry of these species may also be important. Pommer (2003) and Fick (2003) inferred heterogeneous terpene reactions with NO 2 and ozone and radicals on tefl on tubing because observed rate constants were much larger than expected. Similarly, oxidation of terpenes on surface of HVAC heat exchangers was inferred from a larger - than - expected reduction in the concentration between the upstream and downstream locations (Fick et al ., 2005 ). They also observed that the fraction of terpenes oxidized correlated with surface area. Increased ozone deposition rates to surfaces are observed long after terpene primary emissions have ceased (Destail-lats et al ., 2006b ; Singer et al ., 2006a) . The ozone reaction probability of adsorbed terpenes ( δ - limonene and Δ - carene) has been observed to be 1 to 2 orders of mag-nitude greater than that for the gas - phase reactions (Springs and Morrison, 2008 ). Further, different products are generated at surfaces, for α - terpineol (Ham and Wells, 2008 ), than in the gas phase. Together, these studies indicate that adsorbed terpenes promote ozone uptake on indoor surfaces and that overall terpene decom-position is enhanced over homogeneous processes.

Taken as a whole, it is probable that some fraction of ozone uptake and second-ary emission in commercial and residential buildings is due to reactions with soaps, cooking oils, human skin oils, terpenes and other products of human inhabitation. This may partially answer the question posed by Nazaroff, Gadgil

Scheme 13.7 Ozone oxidation of nicotine on cotton and Tefl on (Destaillats et al ., 2006a ).


and Weschler (1993) in their evaluation of ozone deposition velocities on indoor surfaces: why does the area - averaged ozone reactivity of indoor surfaces in most fi eld sites tend to be roughly the same, even though laboratory studies show that individual surfaces are very different and that surfaces age and lose reactivity with time? They suggested that some sort of uniform soiling of surfaces with reactive species may explain this observation. Here we see that human skin oils, cooking oils, and even cleaning all contribute to enhancing the surface reactivity of indoor surfaces. It is probable that these materials also contribute to the enhanced reactiv-ity of HVAC fi lters and surfaces.

13.4.3 Heterogeneous Chemistry: Acid – Base

Acid – base chemistry on a surface has the effect of changing the form of the surface bound species, instead of directly generating new volatile compounds. This can infl uence the overall sorptive capacity of indoor surfaces or even catalyze transfor-mative reactions. Compound sorption infl uences the timing and intensity of expo-sure by temporarily storing these species on indoor surfaces (Tichenor et al ., 1991 ; Won et al ., 2001 ). For example, if a compound adsorbs strongly during an emission event, the peak concentration during the event will be lower than anticipated. However, desorption of those compounds will cause occupants to be exposed over an extended time period.

Acid – base chemistry taking place on surfaces may infl uence adsorption of basic or acidic compounds. It is well established in aqueous systems that pH infl uences the sorption of organic compounds to solids. However, indoor surfaces are not uniformly covered with bulk water, nor are they uniformly dry. Instead surfaces are complex and exhibit a range of molecular water coverage depending on humid-ity, surface composition and morphology. Therefore, the infl uence of acid – base chemistry on sorption is non - obvious. A volatile or semi - volatile organic com-pound that has basic functionality, such as nicotine (Van Loy et al ., 1998 ), may bind strongly to an acidic surface site, or perhaps convert to its conjugate acid salt in acidic surface water. If suffi cient acidity exists, the total surface capacity for that compound can be substantial. Organic acids accumulate in an analogous fashion on basic (alkaline) surfaces. The acidity/basicity of a surface depends on the com-position of that surface, including the substrate (for example, a carpet fi ber), surface amendments (salts, oils and so forth), surface moisture. Therefore, sorp-tion can also be infl uenced by adsorption of overlying gases such as acidic carbon dioxide and basic ammonia.

Evidence of this behavior was reported by Webb, Singer and Nazaroff (2002) . They opened a bottle of an ammonia - based cleaning solution in a room that had been preconditioned with 3200 mg of nicotine. They observed an immediate increase in the gas - phase nicotine concentration that they ascribed to conversion of nonvolatile nicotine salts to semi - volatile free - base nicotine.

Ongwandee et al . studied the adsorption behavior of trimethylamine ( TMA ) on three surfaces: ZrSiO 2 (representative of mineral surfaces) (Ongwandee, Bettinger


and Morrison, 2005 ), painted gypsum board and carpet (Ongwandee and Morrison, 2008 ). They found that ammonia, a base, decreases the surface capacity of all materials. Carbon dioxide, an acid, tends to increase the surface capacity, but only moderately so. This is to be expected if aqueous - phase chemistry is taking place, converting TMA to its less volatile, protonated form. Instead, competitive adsorption, rather than aqueous - phase acid – base chemistry appears to drive most of these dynamics. Changes in surface site acidity may also contribute to this observed behavior.

Combined, these studies demonstrate the importance of evaluating sorption of VOCs under the appropriate conditions, that is, with typical indoor levels of CO 2 , NH 3 , humidity, temperature, and so forth. Since most polar organic compounds are Lewis bases, or are amphoteric, they may also be infl uenced by changes in surface acid sites or the presence of competitive gas - phase species. Destaillats et al. (2006a) note that only the free - base form of nicotine is susceptible to oxida-tion. This may mean that interactions with acid sites on the surface could also suppress the oxidation reaction of this and other amines.


The study of indoor organic chemistry improves our understanding of personal exposure to both reactants and products. At present, our ability to make predic-tions or estimate past exposure is rudimentary. To improve, we need a more comprehensive evaluation of the mechanisms, rates and mediating factors in indoor environments. For example, it is well established that humidity tends to enhance ozone uptake on indoor surfaces, but how does this infl uence product formation? Do CO 2 or NH 3 infl uence transformative product yields as well as infl uencing the sorptive capacity of surfaces? To what extent do occupants contrib-ute to this chemistry through their choice of products, smoking or cooking?

How do we control this chemistry, or use it to our advantage? Controlling this chemistry has been addressed in some product formulations already (e.g., reduc-ing formaldehyde in urea – formaldehyde resin). But as we identify more and more problematic products of indoor chemistry, we will need to take a hard look at manufacturing, construction practices or consumer activities to determine how or if this chemistry can be effectively controlled (Morrison et al ., 2006 ; Uhde and Salthammer, 2007 ). Beyond ‘ control ’ , chemistry may also be used as a tool for directed improvement of indoor environments. Since ammonia has been shown to drive nicotine off surfaces, could we use chemistry to remediate contaminated surfaces indoors? Can ozone be used to selectively destroy compounds on surfaces (such as nicotine), while avoiding the generation of irritating or toxic of by - prod-ucts? These and many other questions remain to challenge us in our efforts to improve indoor environments.

The fi eld of indoor chemistry has moved out of its infancy to a robust adoles-cence. Questions remain, but product yields, rates and mechanisms have been


identifi ed for many reactions. I look forward to many fruitful years of discovery in indoor chemistry.

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Human Responses to Organic Air Pollutants Lars M ø lhave

14

14.1 Introduction

Indoor air pollution is compounds present in the air of buildings in concentrations and for durations which together cause unwanted effects. The pollutants may be inorganic, organic or viable, gases, vapors, or particulate matter ( PM ). In this chapter the indoor environments are in non - industrial buildings where no indus-trial production justifi es the presence of air pollutants. The effects considered are changes of human health and comfort, whereas effects on the building, its appli-ances, plants, animals, etc. are excluded. Three classes of frequently occurring indoor air pollutants are listed; the background for this classifi cation is found in M ø lhave, (2000a) .

1. Gases and vapor s ( GV s) • OCIA s ( organic compounds in indoor air , including irritants, VOCs, SVOCs

etc. see below), also including: � formaldehyde � PAHs � biocides (pesticides etc) � nitrosamines

• radon • NO 2 (also an indicator of combustion products) • CO • CO 2 • SO 2 • Reactive compounds and reaction products (including O 3 etc)

2. Non - viable particulate matter ( suspended particulate matter ( < 10 μ m) ( PM 10 ) • biological debris (dander, metabolic products, and cell fragments, scales,

mites, pets etc) • asbestos • MMMF ( man made mineral fi bers )


327

328 14 Human Responses to Organic Air Pollutants

3. Viable particle s ( VP s) • molds • spores • other microbes.

Volatile organic compound s ( VOC s) are considered in this chapter. The group of organic gases and vapors according to the World Health Organization (WHO, 1982, 1986 ) consists of four classes of pollutants defi ned by boiling point ranges. The class of VOCs contains organic compounds with boiling points from 50 to 100 ° C up to 260 ° C depending on the sampling and analytical techniques (Ber-glund et al. , 1997 ; M ø lhave et al. , 1997 ; WHO, 1989 ). From a toxicological point of view all compounds present as air pollutants are relevant. Toxicologists there-fore request a wider analytical window to be used. In consequence, a Scandinavian working group has recommended that investigations be planned to include the entire range of OCIAs (Andersson et al. , 1997 ).

The OCIA group includes:

• All VOCs according to the previous defi nition by WHO (1982, 1986) • Radicals, ionic species such as salts and surface active compounds • Compounds adsorbed on particulate matter

but excludes:

• Proteins, glucan, etc. • Other organic chemicals with molecular weight greater than 500 – 1000 Da.

Health is a state of complete physical, mental, and social well - being, and not merely the absence of disease or infi rmity (WHO, 1948 ). This implies that the human health effects which are relevant indoors include both adverse effects and changes of well - being associated with exposures to environmental factors. Therefore, the diagnosis of an environmental health effect, includes both identifi cation of a human health problem and an exposure, that is, identifi cation of causality.

To prevent health effects from occurring indoors, guidelines have been estab-lished. A guideline is a set of criteria (i.e., standards for making judgements) specifi cally assembled to indicate threshold levels of a harmful or noxious agent consistent with the maintenance of good health. In the context of indoor air, a criteria standard often describes an exposure – effect relationship between a pollut-ant and the risk of an undesirable effect under specifi ed circumstances. The tra-ditional health effect data used for guideline setting are called the NOAEL ( ‘ No Observed Adverse Effect Level ’ , i.e., the highest exposure level which did not show signifi cant adverse effect of exposure) or the LOAEL (the ‘ Lowest Observed Adverse Exposure Level ’ showing such an effect). The interest in LOAEL and NOAEL is derived from their use as estimates for the threshold of the effect and exposure in question and estimated threshold refers to a specifi c population exposed under specifi ed conditions to specifi ed compounds.

14.2 VOC Exposures Indoors 329

Recommendations are less strict specifi cations of certain actions to be taken under specifi ed circumstances and have the form of advice but no legal consequences or obligations. An ALARA recommendation refl ects the ‘ precaution principle ’ of keeping the exposure ‘ As Low As Reasonably Achievable ’ (see M ø lhave, 2000a ). The Guideline - Recommendation - ALARA refl ects three levels of stringency in guideline setting based on the severity of the effects, their risk or prevalence, the quality of documentation, and the means available for quantifi ca-tion of the effects and exposures. The type of guidance to give in each case depends on many factors including severity, prevalence of the effects, and available health risk data.

Further information about the indoor environment is found in Axelrad et al. , (1995) and Spengler et al. , (2001) . Further discussions of the principles of setting guidelines for indoor air quality ( IAQ ) can be found in Cochet et al. , (2006) .

14.2 VOC Exposures Indoors

In literature, most VOC exposure levels indoors are described as being low, which indicates both that the exposures are low in comparison to occupational threshold limit values ( TLV ), and that health effects are expected to be reversible and mostly unspecifi c, for example, caused by stimulations of the senses such as the common chemical sense. Irreversible adverse effects as known in the occupational environ-ment are generally not expected although adverse effects may occur in rare events (e.g., including hypersensitive persons) or as accidental exposures. The exposures are also described as multi - component exposures consisting of 50 to 200 different VOCs, in concentrations which, considering the large range of toxicity of the individual components, makes it diffi cult to identify any particular toxic compo-nent as being responsible for any effects observed (Berglund et al. , 1991, 1992 ; Bluyssen et al. , 1997 ). Further information on VOC exposures indoors are found in Axelrad et al. , (1995) and Spengler, Samet, and McCarthy, (2001) .

Many compounds may interact and cause unexpected toxic effects which could not be expected from the known toxicity of the individual components. Often one of three principles applies for assessment of the interactions between compounds in a mixed exposure. These are (i) independent action or no interaction between the components; (ii) interaction between the compounds (which in some cases may be dealt with by adding the expected effects of the individual compounds using a weight factor expressing the relative potency of the compounds); and (iii) no assessment is possible because the interaction mechanisms are too complex for normal toxicological evaluation. It should be noted that different types of health effects may require different types of assessments of the interactions and that these assessments have different accuracy. These interactions are the main causes of the slow process of establishing offi cial guidelines of IAQ. Further information is found in Cochet et al. , (2006) .


14.2.1 Health Effects of Indoor Air Pollution

M ø lhave (2000a) has argued for a grouping of the health effects relevant to indoor air into fi ve classes (which will be used in this chapter):

1. Immune effects and other hypersensitivity • asthma • allergy • non - specifi c hypersensitivity

2. Respiratory effects (other than immunological) 3. Cellular effects

• cancer • other cellular effects including reproduction

4. Neurogenic and sensory effects • odor • irritation • neurotoxic symptoms

5. Cardiovascular effects

Generally, the effects or symptoms observed after exposures to low level VOCs do not identify a specifi c causality or a particular causal compound. Symptoms may result from a wide range of causes, such as thermal stresses, mental strains, or from any number of diseases. The mere complexity of such co - variates is a likely reason why many uncontrolled fi eld investigations often fail to demonstrate clear correlations between environmental exposures and symptoms reported with build-ing occupancy. It has been found that perceived IAQ and general well - being is strongly correlated to both irritation and odor (Hudnell et al. , 1993 ; M ø lhave, Bach and Pedersen, 1986 ; Otto et al. , 1990b ). Therefore general evaluations, such as perceived IAQ, general well - being, and the need of more ventilation, are included in this summary.

In the literature, different types of questionnaires have been used to register subjective evaluations. They are all self - administered. Some are paper and pencil questionnaires (Andersson, 1998 ; Fanger, 1988 ; M ø lhave, Bach and Pedersen, 1986 ) and some are administered by personal computers, one for each subject (Kj æ rgaard et al. , 1993 ; Otto et al. , 1992 ). The rating scale differs as some are visual analog scale s ( VAS ) in which the rating is reported using a mark on a line anchored between two extremes while other scales use fi xed phrases (categorical scales) between which the subject has to choose. Despite differences in type of scale and phrasing, in this chapter questions and symptoms are here grouped in the general categories shown in Table 14.1 . For more information on sensory evaluations see Berglund et al. , (1999) .

Several biological models for these unspecifi c responses have been discussed in the literature (e.g., M ø lhave, 1991 ). Table 14.1 summarizes one of the classifi ca-tions or general categories of effects suggested in the literature. They include: perceptions of non - VOC related environmental exposures (such as lighting or


Table 14.1 Ad hoc classifi cation of health effects of organic gases and vapors. Italic marks the symptoms dealt with in this chapter. (General irritation = irritation in eyes, nose, or throat).

Class of symptom Rating of General classes of perceptions

Examples

Perceptions of non - VOC exposures from the environment

Annoyance or intensity Auditory Visual

• Air temperature • Noise.

Perceptions related to VOC exposures

Annoyance, intensity, hedonic mode or strength

Olfaction • Odors a • Perceived IAQ a

Stimulation of the chemical sense

• Perceived irritation of mucosal membranes in eyes, nose, or throat a

• Dryness of eyes, nose, or throat a

• Dryness or humidity of skin a

• Facial skin irritation • Erythema a , b

General evaluations Annoyance, intensity, or strength

• General well - being – Need more ventilation

• Perceived overall air quality • Unspecifi c hypersensitivity a

Perceptions of body symptoms

Annoyance or intensity Symptoms predominantly of infl ammatory origin

• Feeling of watering or runny eyes

• Feeling of blocked or of watering or runny nose

Symptoms predominantly of neurological or stress origin

• Feeling of cough a • Concentration diffi culties • Sleepiness, tiredness or

fatigue a • Headache, sluggishness, or

heavy head a Predominantly other types of symptoms

• Skin temperature • Sweating • Nausea a • Hoarseness a • Facial skin temperature and

humidity

a included in WHO defi nition of SBS (WHO, 1982, 1986 ). b Erythema is the name applied to redness of the skin produced by congestion of the capillaries which may

result from a variety of causes.


thermal environment), perceptions of odors, perceptions of irritation of nose, eyes, upper airways, and general discomfort, or neurological symptoms, such as mental fatigue or headache. The main emphasis is on sensory irritation and related effects. In the literature the term irritation is used for two different types of sensory irritation. One is the perceived sensory irritation caused by an environmental exposure to, for example, pollutants, which stimulate the senses (normally the trigeminal or general chemical sense). The other is often called infl ammatory irritation and refers to stimulation of the senses through mediators released in the body, for example, after exposure of tissue to biological potent pollutants. This classifi cation will be used in this summary. However, it should be noted that no unambiguous scientifi c classifi cation is possible because of the complexity of the responses. The table merely offers an ad hoc vocabulary used in this chapter which will identify symptoms reported to be signifi cantly related to VOC exposure in at least two human exposure experiments. These symptoms are indicated in Table 14.1 in italic. The subjective sensory effects (self - reported) are in focus because they seem to be experienced at lower exposure doses (e.g., time × concentration) than the objectively measurable effects.

Some symptoms have been reported more frequently than other symptoms. One group of such frequent symptoms has been identifi ed as a syndrome of coexisting symptoms related to IAQ. The syndrome is often called the Sick Building Syn-drome and includes the symptoms mentioned above and in Table 14.1 , according to its fi rst defi nition by WHO working groups (WHO, 1982, 1986 ). The syndrome has been further discussed by M ø lhave (1991) . Presently, it is not clear if this SBS syndrome consists of truly correlated symptoms related to one exposure type indoors or refl ects an accumulation of effects of several unrelated exposures indoors. In this chapter, no distinction will be made between immediate acute responses, appearing within the fi rst ten or fi fteen minutes after the onset of exposure, and responses appearing later, even after the exposure has stopped. This is often taken to correspond to the visitors ’ and the occupants ’ responses, respec-tively, and may indicate different time cause of effects.

14.2.2 Indicators of Indoor Air Quality and Health

As long as no generally accepted causality and dose response relations have been established for the IAQ syndromes, the NOAEL or LOAEL for the effects cannot be established. It follows that systematic mitigation is diffi cult as the causality is unclear and the effects of interventions diffi cult to predict. Syndromes may have a function as indicators of a general level of IAQ, and their existence is important for the understanding and planning of rational public health activities.

It follows that it is diffi cult to plan rational public health activities because no accepted causality exists for most of the effects observed indoors and that often no quantifi cation or prediction of the effects of mitigation can be made based on rational criteria. Indicators or substitute measures become alternatives in guide-lines and recommendations for mixtures of pollutants.


Many indicators or substitute measures are in use in public health. Some are exposure indicators. Examples are COhb as an indicator for CO poisoning, nicotine as indicator for tobacco smoke, CO 2 as ventilation indicator, PM 10 as a measure of particulate matter, and Olf/Decipol as an indicator for air pollution with bio - effl uents from occupants. Some are health indicators, such as changed lung function for health effects caused by indoor air environment, and some may have a double function as both exposure and effect indicator (such as SBS).

The general principles of indicators of exposures or effects have not been widely discussed in the literature. Some information is found in WHO, (2001) . From the examples mentioned above it appears that indicator or substitute measures are used when exact measurements of effects or exposures are diffi cult to use for economical, time, or practical constraints. The indicator variable or substitute measure can be any variable with a known (causal or non - causal) association to the variable in question. It should be noted that when causality is established this exposure measure becomes an indicator of the effect and vice versa. In many cases only an association exists and often only within a limited range of exposure situations.

There are several types of exposure indicators. One compound of a mixed expo-sure may be used to represent all compounds of the mixture if the relative ratios between the components are relative stable. (Examples are source indicators used for nicotine/tobacco smoke or CO 2 to indicate bio - effl uents). A balanced sum of the concentrations of all components may be used (assuming no interaction, examples are PM and TVOCs). Biomarkers, for example, in the form of a reaction product related to the metabolism of or reaction to the exposure can be used (examples are enzyme indicating exposure to ETS and COhb indicating exposure to CO). Finally, the physical presence of a source may be used as a crude indicator of exposure (e.g., number of smokers in a room). The use of indicators has many limitations of which an important one is that their use generally decreases the predictive value of exposure assessment. For very toxic indicator measurements can mostly be used only as screening tools (e.g., COhb). For the non - adverse effects, the high prevalence of these effects make a pragmatic simple solution more important than the exact measurements and therefore indicators become a favored tool. This discussion is relevant in relation to VOCs for which several indicators have been suggested (see below).

After the fi rst experiment with low level exposures to a mixture of VOCs had demonstrated that unexpected health and comfort effects may follow from low level exposures (M ø lhave, Bach and Pedersen, 1986 ), the concept of TVOCs was suggested as a measure of a mixture ’ s potency to cause effects. The concept was recently revised (Andersson et al. , 1997 ; M ø lhave and Nielsen, 1992 ; M ø lhave et al. , 1997 ; Seifert, 1999 ). This TVOC concept is still debated. Undoubtedly, it is based on several assumptions that are not yet justifi ed and its general usefulness for the prediction of effects of mixtures other than the M22 - mixture (see below) is undocumented. The appealing aspect of the indicator is that it is easy to measure and use in pollution control and regulation.


Olf and Decipol are other units for indoor pollution. OLF is a unit for emission rate of bio - effl uents and other indoor air pollutants. They refl ect the perceived air pollution caused by emissions from a standard person in thermal comfort. Decipol is a unit for perceived air pollution and one Decipol is the pollution caused by one standard person ventilated by 10 L/s unpolluted air (Fanger, 1988 ).

14.2.3 Classes of Indoor Air Pollutants

In M ø lhave, (2000a) it is argued that three classes of health effects exist in relation to health, IAQ, and guideline setting. They are called priority, secondary, and hypothetical or potential pollutants and health effects related to these pollutants. This chapter mainly deals with secondary health effects.

Priority IAQ pollutants include effects and exposures for which the causality and relationship to the exposure are well established. The basic principle for the toxicological evaluation of these priority causalities is a compound by compound evaluation. It is assumed that criteria exposures are so few and infrequent that interactions between different criteria exposures can be neglected. Often offi cial guidelines or recommendations have been set and quantifi able health criteria or dose – response relations are used in the setting of these health - based guidelines. In some cases insuffi cient knowledge (e.g., on measurements) may prevent the setting of strict guidelines and recommendations may have to be used. Table 14.2 shows some examples. The most important pollutants (called priority pol-lutants) have already been addressed with guidelines or recommendations on national levels. However, an international set of guidelines or recommendations is needed for these pollutants based on general and uniform rules for setting such standards.

Secondary health effects indoors are less adverse, and mostly characterized by being reversible, and with no severe consequences for the occupants. The effects may have high prevalence. For these effects the causality may not be exactly known but an association to IAQ has been documented. Exact quantifi cation may not be possible and indicators or substitute measures may have to be used. For secondary effects strict guidelines cannot be set and a strict compound - by - compound evalu-ation is not possible (or needed).

Recommendations are the preferred tools and an ALARA procedure may be used. Table 14.2 shows some examples. Secondary pollutants may be grouped into a few classes based on considerations such as similar biological target effects or biological mechanisms, similar principles or models for toxicological evaluation, similar procedure for health data, similar analytical procedures, availability of measures of effects or exposures, or availability of health data. In this context the TVOC indicator is discussed as explained below. Some of these less adverse pol-lutants (here called secondary pollutants) have also been addressed with guidelines but the present knowledge on quantifi cation of exposure, health data etc. prohibit the setting of strict offi cial guidelines. Major research is needed on many of these secondary causalities before international recommendations or guidelines can be


Table 14.2 Summarizes of IAQ relevant pollutants and levels of guidelines (M ø lhave, 2000a ).

Type of causality Causality Level of required regulation

Priority pollutants Allergy & molds and other microbes. Includes rhino conjunctivitis.

Recommendation or ALARA

Allergy & biological debris Guidelines Cancer & OGV Recommendation/ALARA Cancer & radon Guidelines Cancer & asbestos ALARA Respiratory effects & NO 2 Guidelines Respiratory effects & molds,

other microbes and spores Recommendation

Irritation & formaldehyde Guideline Cardiovascular effects & CO Guideline

Secondary pollutants Irritation and MMMF Recommendation Respiratory effects & PM Recommendation Odors & gases and vapors Recommendation Irritation & OGV Recommendations/ALARA Irritation and CO 2 Recommendation

Hypothetical relevant pollutants Asthma & OGV ALARA Cancer & PAH ALARA Cancer & biocides ALARA Cancer & formaldehyde I Irritation & reaction products ALARA Irritation & endotoxins or Glucan ALARA Neurotoxicity & biocides ALARA

established. In the interim period, a precaution principle should lead to an ALARA principle for these secondary causalities.

Hypothetical or potential causalities include postulated causalities which have not yet been proven to be related to IAQ. An example is multiple chemical sensi-tivity ( MCS ). Often effects or exposures cannot be quantifi ed, not even by indica-tors. Toxicological evaluations can only deal with effects and exposures which are quantifi able. Therefore, these undocumented causalities must be defi ned as hypo-thetical or potential causalities. Even if some of these effects may be considered adverse or have high prevalence, the present level of knowledge prohibits an offi -cial rational preventive action and no offi cial guidelines or recommendations can be established. It should be noted that all new types of IAQ problems start as hypothetical. This group includes many health effects which are potentially impor-tant to human health and thus defi nes the research forefront in IAQ sciences. Table 14.2 shows some examples. Several potential or hypothetical causalities are discussed in the scientifi c literature. Their relevance to health indoors is unknown and no offi cial recommendations or guidance can be established. Mitigation of indoor climate problems becomes a matter of ad hoc solutions and a major


research effort is needed to establish facts for these hypothetical effect exposure relations. In the meantime, a general ALARA precaution principle should be applied assisted with ‘ trial and error ’ - based mitigation.

From the WHO statements on rights to a healthy indoor air (M ø lhave and Krzyzanowski, 2003 ; WHO, 2000 ) it follows that national and international orga-nizations have an obligation to establish criteria for good IAQ. Such health guide-lines are needed as targets for associated guidance, for example, on ventilation, building codes, and consumer safety. In principle, no preventive action should be made before a set of criteria has been established on which its economical health or other consequences can be judged, the success of the activity can be defi ned (success criteria), prioritization of preventive actions can be made, and which defi nes when rational preventive actions are possible.

Any set of IAQ guidelines or recommendations intended to cover all relevant pollutants including VOCs must be based on the same quality criteria and must deal with: all relevant causalities (exposures and effects); interactions in a multi - factorial exposure matrix, using the same priorities for IAQ relevance and accept-ability, evaluation procedures, safety levels; the effects modifying action of non - chemical factors (e.g., temperature, noise etc) on human sensitivity, chemical interactions in the air; and protect the entire population against health effects caused by a specifi ed and defi ned exposure scenario. As long as such guidelines are missing, unnecessary exposures to pollutants must be avoided. This translates into an ALARA principle for all indoor air pollution exposures until an offi cial guideline or recommendation has been set.

14.2.4 The TVOC Indicator

In a regime of ‘ not knowing what to do ’ the TVOC indicator was born. The TVOC ’ s mass/m 3 was suggested as an indicator of exposure to a mixture of VOCs and of its potency to cause effects reviewed by M ø lhave (2000b, 2001) . From the begin-ning it was underlined that this TVOC indicator is intended as an ad hoc tool meant for screening and for sensory irritation (i.e., non - adverse secondary causali-ties as described above). The indicator is based on the assumption that additivity can be used to calculate a summation index which indicates low risk of effects when below one.

Subsequently, it has been argued that the individual compounds must be bal-anced according to their toxicity (M ø lhave and Nielsen, 1992 ). As a consequence, a balanced evaluation based on weight factors called LCI has been suggested (Bluyssen et al. , 1997 ; Berglund et al. , 1997 ).

The TVOC indicator is and has been widely misused. The indicator is not an offi cial recommendation or guideline and no defi nitive conclusions should be made based on this indicator alone. It is a screening tool needed for exposure evaluation, source identifi cation, and IAQ evaluation. However, as stated by the NORDVOC committee (Andersson et al. , 1997 ) there is insuffi cient evidence to either accept or reject this hypothetical ad hoc tool. It fi nds some analogy in dB(A),


TSP (mass), and total PAN. However, the best argument for its use is still that few alternatives exist that are equally easy to use in practice. With many reservations some offi cial organizations therefore point to the use of TVOC as an exposure indicator (Seifert, 1999 ; Bluyssen et al. , 1997 ; Berglund et al. , 1997 ; ASHRAE, 2004 ). The TVOC concept is based on several assumptions and its usefulness for prediction of health effects of mixtures is undocumented. The concept does not include the possibility of interactions between the many compounds in indoor air and no toxicological arguments for exclusion of some and inclusion of other toxi-cologically relevant organic vapors and gases. In addition it must be remembered that different VOCs have different toxicity and that different laboratories use dif-ferent measuring procedures. There are no true standardized procedures for TVOC measures.

It follows that TVOC is diffi cult to use for normal regulatory risk assessment; the scientifi c basis for this is just too small and no D – R relations have been estab-lished. TVOC at this point should only be used for screening and not for defi nitive conclusions. In addition, TVOC should only be associated with sensory irritation and only if there are substantial indications that VOC is a problem. In each specifi c case, if unusual compounds and concentrations are identifi ed, the use of TVOC should be stopped. If TVOC is in the mg/m 3 range, additional alternative methods should be used to draw any conclusions.

14.3 Summary of Experimental Evidence of Health Effects of VOC Exposure

VOCs were fi rst suggested as a potential cause for some of the indoor air symp-toms reported during the 1980s (M ø lhave, 1982, 1986 ). However, fi eld investiga-tions in buildings where the occupants complained about reduced IAQ and discomfort often failed to demonstrate the reason for these symptoms. However, the symptoms were similar to those known to follow from low level exposure to VOCs in occupational environments.

Since then much research has been done on the possible contributions of VOCs to indoor air problems. Results of human exposure experiments were reviewed by M ø lhave (2000b, 2001) . Several controlled experiments were performed in labora-tories around the world in which human responses to VOCs known as indoor air pollutants were investigated to test among other things, to what extent low level exposures to such VOCs might contribute to the prevalence of complaints in non - industrial buildings. This chapter summarizes some of the fi ndings in 12 of these controlled experiments.

14.3.1 Symptoms Relevant to VOC s

Table 14.3 shows the 12 studies and Table 14.4 summarizes the fi ndings of the 12 studies. The table refers to SBS symptoms related to VOC exposures that have

338

14 Hum

an Responses to O

rganic Air Pollutants

Table 14.3 Summary of the 12 VOC exposure experiments (M ø lhave, 2001 ).

Study no References Subjects type Exposure type Exposure mg/m 3

Exposure duration

Notes

1 M ø lhave, Bach and Pedersen, 1986 ; Thygesen et al. , 1987

62 Healthy adults, 18 – 64

M22 0, 5, 25 2.75 h; 165 min

Subjects who had never experienced SBS symptoms were excluded.

2 M ø lhave, Jensen and Larsen, 1991

25 Healthy adults, 16 – 64

M22 0, 1, 3, 8, 25 50 min Men and women, smokers and non - smokers.

3 Hudnell et al. (1992) ; Otto et al. (1990a,b, 1992)

76 Males, 18 – 39 M22 0, 25 2.75 h Time cause study, subjective and neurobehavioral testing.

4 Johnsen et al. , 1991 ; Wolkoff et al. , 1991

20 Asthmatics, 5 healthy

Materials emission 0, 1.1 – 2.0 a 6 h Not balanced design. Other non - VOC exposures added.

5 Wolkoff et al. , 1992 30 Healthy females Emissions from offi ce machines

0.071 – 0.087 a 6 h Not balanced design. Other non - VOC exposures added. Reported as Olfs.

6 Otto et al. , 1993 26 Males, 15 Females

M22 0, 25 4 h Gender differences. Symptoms and neurobehavioral tests.

14.3 Sum

mary of Experim

ental Evidence of Health Effects of V

OC

Exposure 339

Study no References Subjects type Exposure type Exposure mg/m 3

Exposure duration

Notes

7 Prah et al. , 1993 20 Healthy males, 18 – 35

M22 0, 12, 24 4 h Effects on respiration. Also mouse irritation testing.

8 M ø lhave et al. , 1993b ; Kj æ rgaard and M ø lhave, 1994

30 Healthy males and females, 19 – 60

M22 and subsets 0, 1.7, 5, 15 1 h Reported symptoms and irritation of ear, nose, throat.

9 Kj æ rgaard, M ø lhave and Pedersen, 1991

21 Healthy 14 SBS - sensitive

M22 0, 25 2 h Risk group investigation. Reported odors and irritation.

10 M ø lhave et al. , 1993a ; Kj æ rgaard et al. 1993

10 Healthy M22 0, 10 1 h Thermal interaction (3 temp settings 18, 22, and 26 ° C.) Objective measures symptom reporting and irritation of ear, nose, throat.

11 Kj æ rgaard et al. , 1995a,b 18 Healthy 18 Sensitive/ hay fever patients

M22 0, 20 4 h Test of asthmatics/hay fever patients, non - smokers, subjective and objective measures.

12 Hudnell, Otto and House, 1993

46 Healthy males M22 0, 6, 12, 24 4 h Time course study with symptom rating.

a Other exposure variables included. MIX 22 = Mixture of 22 VOCs.


Table 14.4 Summary of estimated LOAEL and NOAEL of sensory effects observed more than once in the 12 VOC experiments.

Effect Lowest LOAEL mg/m 3

Highest NOAEL mg/m 3

Number of observations

Remarks

Perceived air quality 1.7 5 12 Depending on VOC mixture & temp

Odor intensity 1.7 1 12

Irritation of: • eye • nose • throat

8 1.7 5

3 5 25

12

Feeling of dryness of eye, nose, throat

No data 25 1

Feeling of skin humidity, dryness or irritation

10 20 3

Perceived watering or runny eyes

10 (24) 2 Assuming no report = no effect at highest exposure

Blocked, watering or runny nose

20 No data 1

Additional ventilation needed

1.7 3 5

Feeling of cough 1.7 25 3

General well - being 10 8 4

Headache or heavy head 20 No data 5

Concentration diffi culties 20 12 5

Feeling of sleepiness or tiredness

25 25 3

been observed in at least two of the reported studies. The studies cover exposure ranges from 0.087 mg/m 3 to 25 mg/m 3 . However, only those in the range from 1 to 25 mg/m 3 are conclusive with respect to the effects of VOC exposures indoors. The non - conclusive studies illustrate the importance of co - factors and underline that the following only can be used as estimates of the upper level of thresholds of the effects of mixtures of VOC.


The experiments dealt with exposures to mixtures of selected VOCs, individual VOC compounds, emissions from building materials, or from offi ce machines. One specifi c mixture of VOCs (M22) has frequently been used (M ø lhave et al. , 1986, 1991 ). In the experiments, the relative concentrations of the constituents were constant as indicated in Table 14.3 . Only the total concentration varied in the experiments based on this type of exposure.

Not all experiments can supply estimates of both NOAEL and LOAEL or even one of these toxicological key data. If only one exposure level is used, only one of the estimates can be made. Even if two exposure concentrations were used, the resulting data may turn out to be too few or the variance too big to allow a detailed statistical analyses. In such cases, only differences between exposure and non - exposure can be tested. Also often only positive fi ndings are reported. Therefore, is it not possible to estimate NOAEL in such cases.

Considering that at least two studies should have shown effects and that the thresholds should be relevant for the low level indoor exposure range (e.g., below 10 mg/m 3 ) then Table 14.5 shows a list of symptoms and thresholds (mg/m 3 ) in relation to M22. Assuming that M22 is a best case, then at exposures below about 2 mg/m 3 (TVOC) perceived air quality, odor intensity, irritation of eyes or nose, additional ventilation needed, and cough, are expected to be the most sensitive indicators of VOC exposures. However, it must be kept in mind that these symp-toms are unspecifi c and may have many other causes. Therefore, the presence or absence of these symptoms cannot infer that VOC concentrations indoors are the responsible agent, only indicate the possibility.

Table 14.5 Symptoms following VOC exposures at low levels in two or more of the reviewed papers.

Symptom Range of thresholds indicated (mg/m 3 )

Number of reported cases

Perceived air quality 1.7 – 5 12 Odor intensity 1.0 – 1.7 12 Irritation of eyes or nose 1.7 – 8 12 Additional ventilation needed 1.7 – 3 5 Feeling of cough 1.7 – 25 3 Irritation of throat 5 – 25 12 General well - being 8 – 10 4 Feeling of skin humidity 10 – 25 3 Headache 20 5 Concentration diffi culties 12 – 20 5 Feeling of sleepiness 25 3


14.3.2 Effect of Exposure Types

One study showed that the thresholds of effects were above a TVOC level of 1.7 mg/m 3 . Thus the fi ndings in two other studies of effects during exposures at much lower VOC exposure levels clearly demonstrates the diffi culties in using TVOC as an exposure indicator in relation to VOC indoors. This observation is in line with the recommendations of several international consensus groups (Anders-son et al. , 1997 ; Bluyssen et al. , 1997 ; M ø lhave et al. , 1997 ; Seifert, 1999 ). Two studies have specifi cally addressed the differences in potency of different VOC exposures. The studies are not conclusive on this matter. They indicate differences in potency but the relative size of the differences are not presented. However, the authors state that for practical purposes where the exposure can be expected to exceed, the thresholds of effects the differences between different types of expo-sures at the same TVOC level seem to be small. This supports the use of the TVOC indicator.

14.4 Conclusions

From the discussions above it appears that the most potent pollutants (called prior-ity pollutants) have already been addressed with individual guidelines or recom-mendations. However, an international set of guidelines or recommendations for IAQ is needed for these pollutants based on general and uniform rules for setting such standards.

Some of the less adverse pollutants (called secondary pollutants) have been addressed with guidelines but the present knowledge on quantifi cation of expo-sure, health data, etc makes it diffi cult to set offi cial guidelines. Major research is needed on these secondary pollutants before recommendations or guidelines can be established. In the interim period, a precautionary approach should lead to an ALARA principle for these secondary causalities.

Several potential or hypothetical relevant pollutants are discussed in the scien-tifi c literature. Their relevance to health indoors is unknown and no offi cial recom-mendations or guidance can be established. Mitigation of indoor climate problems become a matter of ad hoc solution and a major research effort is needed to estab-lish facts for these hypothetical effect exposure relations. In the meantime, an ALARA principle should be adopted for these exposures.

Indicators can be and are used for both guidelines and recommendations. However, no data support the use of any known VOC as an indicator for the pres-ence of other pollutants. There is a need for research on indicators for the second-ary exposure – effects relationship.

TVOC is an indicator for the presence of VOCs indoors. The TVOC indicator can be used in relation to exposure characterization and source identifi cation but for VOCs only. In risk assessment, the TVOC indicator can only be used as a

References 343

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WHO ( 1989 ) WHO indoor air quality: organic pollutants (eds V.C. Armstrong, B. Berglund, M.A. Berry, J.S.M. Boleij, I. Farkas, R.H. Jungers, M.D. Lebowitz, J. Lewtas, T. Lindvall, L. M ø lhave, D.J. Moschandreas, B. Seifert, J.A. Stolwijk, H. Kn ö ppel and M.J. Suess) Report of the Berlin meeting 1987, Report Vol. 111 , WHO Regional offi ce for Europe . Copenhagen. Denmark , pp. 1 – 69 .

WHO ( 2000 ) The right to healthy indoor air , report of a WHO meeting. European Health 21 targets 10, 13 (eds L. M ø lhave, N. Boschi, M. Krzyzanowski, K. Aas, J.V. Bakke, V. Bencko, M. Chuchkova, C. Cochet, I. Farkas, A. Garriga - Trillo, S. Kakari, P. Kalliokoski, A. Kessel, H. Levin, T. Lindvall, J. McLaughlin, I. Mocsy, G. Muzi, A. Pickering, B. Seifert, K. Slotova, C.L. Soskolne and M. Tallacchini), WHO Regional offi ce for Europe, Copenhagen, Denmark.

WHO ( 2001 ) Biomarkers in Risk Assessment: Validity and Validation , EHC222, WHO Head Quarter, Geneva, Switzerland.

Wolkoff , P. , Nielsen , G.D. , Hansen , L.F. , Albrechtsen , O. , Johnsen , C.R. , Heinig , J.H. , Franck , C. and Nielsen , P.A. ( 1991 ) A study


of human reactions to emissions from building materials in climate chambers. Part II: VOC measurements, mouse bioassay, and Decipol evaluation on the 1 – 2 mg/m 3 TVOC range . Indoor Air , 1 ( 4 ), 377 – 88 .

Wolkoff , P. , Johnsen , C.R. , Franck , C. , Wilhardt , P. and Albrechtsen , O. ( 1992 ) A study of human reactions to offi ce machines in a climatic chamber . Journal of Exposure Analysis and Environmental Epidemiology , 1 , 71 – 96 .


Emission Studies Part Four

Volatile Organic Ingredients in Household and Consumer Products Godwin A. Ayoko

15

15.1 Introduction

People living in modern industrial societies inevitably come into contact with a variety of chemical products. The spectrum of chemicals used for hobbies, clean-ing, cosmetics, professional activities etc is enormous (see Figure 15.1 ), and the enticement for their use is prompted by aggressive advertising conducted by manufacturers and marketers of consumer products (Selinger, 1998 ; Emsley, 1997 ; Wolkoff et al ., 1998 ). Many of these products contain pesticides, solvents, reactive monomers and other hazardous materials which are released during use and which may affect human comfort and well - being (Ott and Roberts, 1998 ). Thus, the pollution of indoor air by VOCs has been considered as a hygienic problem that has necessitated systematic investigations since the early 1980s.

Nowadays, numerous organic pollutants are monitored in indoor air (Krause et al ., 1987 ; Brown et al ., 1994 ; Ayoko, 2004 ; Salthammer, 1994 ; Pluschke, 1996, 2004 ), on house dust (Wolkoff and Wilkins, 1994 ; Wilkins et al ., 1993 ; Butte, 2003, 2004 ) and other adsorptive media. The analytical results from such investigations are usually evaluated as immission and emission. Investigation of indoor air is a typical measure of immission and its result is usually expressed in μ g/m 3 , which denotes the average concentration of a certain substance in indoor air over the sampling period. In contrast, an emission measurement, for example, in test chambers or emission cells (see also Chapters 5 and 6 ), defi nes the release from a defi ned emission source as a function of time. The result of such measurements are generally expressed as area specifi c emission rate 1) ( SER a ) in μ g/(m 2 h) or unit specifi c emission rate ( SER u ) in μ g/h. Alternatively, when the air exchange and the loading factor are taken into account, results of chamber measurements can also be expressed as test chamber concentration in μ g/m 3 .

The determination of the emission versus time behavior under living conditions can be modeled by the use of test chambers or emission cells (see Chapter 6 ).


349

1) The term ‘ area specifi c emission rate ’ is used in parallel with the term ‘ emission factor ’ .

350 15 Volatile Organic Ingredients in Household and Consumer Products

Decay curves whose shape can be described by multi - exponential functions under constant climatic conditions are generally observed. In contrast, the emission properties of substances from household products are related to human activities. Therefore, emission properties of household products depend on how the prod-ucts are used. For example, moth crystals and toilet deodorizers are designed for continuous use. Consequently, they emit volatile components at a constant emis-sion rate. On the other hand, spontaneous release of VOC from sprays, waxes, liquid cleaners and other detergents leads to short - time high concentrations, which decay rapidly.

This chapter summarizes qualitative and quantitative information on emission behavior from literature and provides an overview of the potential for exposure to VOCs during the use of household and consumer products.

15.2 Literature Survey

The spectrum of human activities that is related to VOC emissions is wide. Wolkoff (1995) pointed out that the variety of VOC classes associated with these activities is even greater than that associated with emissions from building materi-als. His publication summarized 17 selected examples of human activities and related VOC emissions from 37 references. Furthermore, Wolkoff (1995) referred

Figure 15.1 Typical activities in dwellings and related applications of household products.

15.3 Product Classes 351

to Rodes, Kamens and Wiener (1991) who stated that human activity - related sources may be regarded as point sources and are closely located to occupants, forming spatial and temporal concentration patterns. A total of 1159 different household products were analyzed by Sack et al . (1992) for 31 VOCs and the results have been compiled into a database. Brown et al . (1994) present SER a s for 13 dif-ferent types of wet household products. As many as 54 VOCs were identifi ed from various cook stoves used in China (Tsai et al ., 2003 ). Several ethylene - based glycol ethers, terpene hydrocarbons, terpene alcohols and other unsaturated compounds were detected in cleaning products and air fresheners (Singer et al ., 2006 ) while more than 200 VOCs were detected after cleaning (Wolkoff et al ., 1998 ). Kn ö ppel and Schauenburg (1987, 1989) identifi ed more than 80 VOCs when investigating eight waxes and two detergents. More data and references are available in Maroni, Seifert and Lindvall (1995) ; Pluschke (1996) ; Witthauer, Horn and Bischof (1993) ; Raaf (1992) ; Selinger (1998) .

15.3 Product Classes

Different classes of products are discussed in the following sections. Table 15.1 shows VOCs, which have been identifi ed by emission testing or material analysis and are representative for the assigned products. In Table 15.2 SER a s and SER u s are summarized for selected products and compounds in order to give an overview of the types VOCs emitted by different products. As mentioned earlier, the emis-sion behavior of household products strongly depends on the type of application and on the test conditions. For a detailed understanding of this behavior,it is necessary to consult the cited references.

15.3.1 Newspaper and Journals

A variety of chemicals such as resins, optical brighteners, fi llers, dye pigments, solvents and others compounds are required for the manufacture of paper prod-ucts like newspaper, journals and magazines (Baumann and Herberg - Liedtke, 1993 ). For example, solvent - based systems, which may contain ketones, esters, aliphatic hydrocarbons and aromatic hydrocarbons are widely used for making printing inks. However, water - based dispersions, whose application have grown continuously over the past years, contain small quantities of solvents, mostly alcohols and glycols.

Braungart et al . (1997) have studied VOC emissions from newspapers and news magazines in a desiccator purged with nitrogen. The main components detected in the chamber air were toluene, hexanal, aliphatic hydrocarbons and α - pinene and the unit specifi c emission rates SER u for the sum of detected VOCs one day after starting the tests were reported to be 128 μ g/h and 143 μ g/h for newspaper and news magazine, respectively. Although the highest emission rate of a single


Table 15.1 Volatile ingredients of household and consumer products. See original references for analytical details.

Product Ingredients Lit.

Pest control agent cis - /trans - permethrin, deltamethrin, pyrethrum, piperonylbutoxid

1

Newspaper, Newspaper journal

Toluene, o - , m - , p - xylene, α - / β - pinene, limonene, aliphatic hydrocarbons (C10 - C20)

2

Electric shaver 1 - ethoxy - 2 - propanol,cyclohexanone, toluene, naphthalene, methylnaphthalenes (isomers), cyclohexanone, C3 - /C4 - benzenes, aliphatic hydrocarbons, BHT

2

Portable CD - player Toluene, C2 - /C4 - benzenes, cyclohexanone, 2 - ethylhexanol

2

Furniture polish (spray) Aliphatic hydrocarbons (C7/C8), branched alkanes

3

Carpet cleanser (spray) 1 - methoxy - 2 - propanol, 2 - methoxy - 1 - propanol

3

Floor wax paste Limonene, α - / β - pinene, Δ 3 - carene 3,6,12

Cleanser/detergent Ethanol, 2 - propanol, 2 - methyl - 1 - propanol, 3 - butenylpropylether, 3 - methylbutyl - acetate, 2 - propanol, limonene, tridecane, 2 - butoxyethanol

4,5,6, 11, 12, 21

Liquid wax Acetone, 2 - butanone, toluene, 1 - hexanol, 3,7 - dimethyl - 1,7 - octanediol, α - pinene, linalool, camphor, linalylacetate

4,5,6

Scientifi c journal Formaldehyde, toluene, pentanal, hexanal, nonanal, aliphatic hydrocarbons (C8 - C18)

7

Insect - spray Aliphatic hydrocarbons (C4 - C12), branched alkanes, cycloalkanes, tetramethrin, D - phenothrin, piperonylbutoxid

8

Furniture beetle agent Aliphatic hydrocarbons (C10 - C14),branched alkanes, cycloalkanes, acetone, dipropyleneglycolmonomethylether, cyfl uthrin

8



Air freshener α - pinene, limonene, myrcene, linalool, octanal, nonanal, α - terpineol, decanal, ocimene, linalylacetate, styrene, chloromethane, propylene, o - dichlorobenzene

8, 14, 15

Paint remover Acetone, 1 - methoxy - 2 - propanol, 2 - (2 - methoxyethoxy) - ethanol, 1 - methyl - 2 - pyrrolidone, C2 - C6 - benzenes

8

Toilet deodorizer Myrcene, limonene, terpinene, terpinolene 8

Adhesive (liquid, high stability) Branched alkanes, cycloalkanes, ethylacetate, 1,1 - diethoxyethane, n - butylacetate, C2 - benzenes

8

Adhesive (liquid, all purpose) acetone, methylacetate, ethylacetate, 2 - butanone, 1 - ethoxy - 1 - methoxyethane, 1,1 - diethoxyethane, 2,2 - diethoxypropane, n - butylacetate

8

Adhesive remover 1 - propanol, 1 - propoxy - 2 - propanol, 2 - propanol, tripropyleneglycole

8

Oven cleaner (spray) Aliphatic hydrocarbons(C4), n - methyl - 2 - pyrrolidone, propyleneglycol

8

Specialized cleaner Acetone, ethanol, 2 - butanone, 3 - methyl - 2 - butanone, 2,2 - diethoxypropane

8

Leather polish 2 - propanol, aliphatic hydrocarbons, branched alkanes

8

Furniture polish (spray) Branched alkanes, cycloalkanes, 1 - butanol - 3 - methyl - acetate, tetrahydronaphthalene

8

Furniture polish (liquid) 1 - methoxy - 2 - propanol, n - butylacetate, 2 - propanol, aliphatic hydrocarbons, branched alkanes, cycloalkanes, 1 - butanol - 3 - methylacetate

8

Shoe polish Aliphatic hydrocarbons, branched alkanes, cycloalkanes, C2 - C4 - benzenes

8

Hair lacquer (spray) ethanol, ethylacetate, branched alkanes, limonene, n - octylether, 1,1,1 - trichloroethane

8, 15




Furniture polish aliphatic hydrocarbons, ethylbenzene, limonene

9

Schoolbook Acetone, cyclohexanone, C3 - benzenes, 1 - butanol - 3 - methoxyacetate, vinylacetate

8

Room freshener Aliphatic hydrocarbons (C9 - C11), branched alkanes, limonene, substituted aromatic hydrocarbons

9

Aircraft disinection Pyrethrins, resmethrin, D - phenothrin, permethrin 13

Hair spray 1,1,1 - trichloroethane 15

Glue Chloroform, ethylbenzene, m - xylene 15

Wool Ethanol, acetone, 2 - methylpentane, 2 - butanone, hexane, methylcylcopentane, pentanal, toluene, methanthiol, DMSO, DMSO2

16

Computers Phenol, toluene, 2 - ethylhexanol, formaldehyde, styrene

17

Photocopiers Benzene, toluene, ethylbenzene, xylene, styrene 18, 19, 20

References: (1) Berger - Prei ß et al ., 1997 ; (2) Braungart et al ., 1997 ; (3) Colombo et al. , 1990 ; (4) Kn ö ppel and Schauenburg, 1987 ; (5) Kn ö ppel and Schauenburg, 1989 ; (6) Person et al ., 1990 ; (7) Salthammer et al ., 1997 ; (8) This work; (9) Tichenor and Mason, 1988 ; (10) Clausen et al ., 1998 ; (11) Wolkoff et al ., 1998 ; (12) Zhu et al. , 2001 ; (13) Berger - Prei ß et al ., 2006 ; (14) Singer et al ., 2006 ; (15) van Winkle and Scheff, 2001; (16) Lisovac and Shooter, 2003 ; Bako - Biro et al ., 2004 ; (18) Lee et al. , 2006 ; (19) Stefaniak et al ., 2000 ; (20) Leovic et al ., 1998 ; (21) Zhu et al ., 2005 .


component (toluene, news magazine), was found to be 88 μ g/h, the study reported no details about test conditions (temperature, humidity, air exchange).

It is a known fact that many freshly printed books and journals have strong odor. In a particularly severe case, school children complained that they were irritated when they touched a schoolbook with a PVC cover. To fi nd out the reasons, the book was conditioned for 24 h and put into a 23.5 - l chamber with open pages at T = 23 ° C, RH = 45% and N = 1.0 h − 1 . Sampling was performed after 24 h and 48 h. As shown in Table 15.3 , very high emissions of aromatic hydrocarbons, glycols and other compounds were observed. In addition, a chamber test carried out on issue 4/96 of Indoor Air journal showed that it had an unpleasant smell. The main components detected in the chamber air were formaldehyde (335 μ g/m 3 ), hexanal (15 μ g/m 3 ), toluene (123 μ g/m 3 ) and aliphatic hydrocarbons ( > C10) (100 – 150 μ g/m 3 ). Higher aliphatic aldehydes (C7 - C11) appeared in concentrations < 10 μ g/m 3 .


Table 15.2 Area specifi c ( SER a ) and unit specifi c ( SER u ) emission rates for different household and consumer products. See original references for analytical details.

Agent Compound SER a μ g/(m 2 h) SER u μ g/h Lit.

Toilet deodorizer TVOC 1.3 10 6 – 3.7 10 6 5 Room fresheners TVOC 1.6 10 5 – 2.0 10 6 5 Waxes TVOC 1.0 10 6 – 9.4 10 7 5 Adhesives (solvent) TVOC 5.1 10 6 – 1.7 10 7 5 Adhesives (water) TVOC 1.0 10 4 – 2.1 10 6 5 Floor cleaners TVOC 1.0 10 4 – 1.5 10 5 5 Moth crystals TVOC 1.4 10 7 7 Wax (fl oor) TVOC 2.0 10 7 7 Cleaning agents/ Insecticides

1,1,1 - trichloroethane 2.2 10 3 6,8 carbon tetrachloride 4.3 10 3 6,8 p - dichlorobenzene 26.4 6,8 m - dichlorobenzene 33.6 6,8

Wax paste (furniture) TVOC 2.6 10 8 3,4 Liquid wax (fl oor) TVOC 9.6 10 7 3,4 Wax paste (leather) TVOC 3.3 10 6 3,4 Cleanser/detergent TVOC 1.1 10 6 3,4 Liquid wax (marble) TVOC 4.8 10 5 3,4 Liquid wax (wood) TVOC 3.0 10 5 3,4 Detergent TVOC 2.4 10 5 3,4 Liquid wax (marble) TVOC 1.8 10 5 3,4 Liquid wax (fl oor) TVOC 1.8 10 5 3,4 Liquid wax (ceramic) TVOC 1.2 10 5 3,4 Desinfectant TVOC

bornyl acetate 3.5 10 4 2

2.9 10 4 2

Detergent (fl oor) TVOC mentol

2.2 10 3 2 1.3 10 3 2

Spray cleanser (carpet) 1 - methoxy - 2 - propanol 5.0 10 4 2 Spray polish (furniture) TVOC

n - octane 2.7 10 4 2

3.7 10 3 2

Wax paste (fl oor) TVOC α - pinene

Answering machine TVOC 51 1 Computer mouse TVOC 26 1 Telephone TVOC 34 1 Electric shaver TVOC 207 1 CD - player TVOC 61 1

References: (1) Braungart et al ., 1997 ; (2) Colombo et al ., 1990 ; (3) Kn ö ppel and Schauenburg, 1987 ; (4) Kn ö ppel and Schauenburg, 1989 ; (5) Person et al ., 1990 ; (6) Sheldon et al ., 1988a ; (7) Sparks et al ., 1990 ; (8) Wallace et al ., 1987 .


Table 15.3 VOC emissions from a school book with PVC cover. Chamber concentrations in a 23.5 l chamber 24 h after loading (T = 23 ° C, r.h. = 45%, N = 1.0 h − 1 ).

Compound Concentration ( μ g/m 3 )

Acetone 67 Cyclohexanone 412 Σ C3 - benzenes a 1643 1 - butanol - 3 - methoxyacetate 1272 2 - butoxyethylacetate 644 Vinylacetate 72

a calculated in equivalents of 1,3,5 - trimethylbenzene.

These results demonstrated that books and journals may release signifi cant amounts of VOCs and must be regarded as typical point sources that are located close to occupants (Salthammer, 1999 ). Consequently, chamber concentrations are presented instead of unit specifi c emission rates.

15.3.2 Insecticides

In the indoor environment, many types of products such as crystals, sprays and liquids are applied for active and preventative protection of insects. Insect sprays are particularly popular because they are easy to handle and can be combined with air fresheners. Commonly, the amount of active agents in these products is well below 2%. For example, a commercially available insecticide for indoor use may contain 0.25% tetramethrin, 0.05% D - phenothrin and 1% of the synergist pipero-nyl butoxide. Pyrethroids are also used as active agents in liquid products against furniture beetle. In addition, materials containing natural fi bers are often equipped with synthetic pyrethroids as a precaution. Apart from the active ingredients, vola-tile components such as acetone, aliphatic hydrocarbons, cycloalkanes, branched alkanes C3 - benzenes and dipropylene glycol monomethyl ether are usually present in insecticides.

Berger - Prei ß et al . (1997) applied different pest control agents for professional use in a model house. The concentrations of active substances were monitored in the gas phase, on suspended particles, on house dust and on furniture surfaces over a period of 24 months. 2 μ g/m 3 and 40 μ g/m 3 of permethrin and deltamethrin respectively were detected on suspended particles immediately after the applica-tion while in the house dust, the initial concentrations of permethrin and delta-methrin were 50 mg/kg and 150 – 800 mg/kg, respectively.

Tichenor and Mason (1988) have shown that m - dichlorobenzene and p - dichlorobenzene are common ingredients of moth crystals. Thus, area specifi c emission rates up to 1.4 × 10 7 μ g/(m 2 h) of these compounds were measured by Sparks et al . (1990) . Similarly, Sheldon et al . (1988a, 1988b) and Wallace et al .


(1987) found the SERs of 33.6 μ g/(m 2 h) and 26.4 μ g/(m 2 h) for m - dichlorobenzene and p - dichlorobenzene, respectively. On the other hand, several volatile compo-nents were detected as ‘ inert ingredients ’ of Foray 48B, an insecticide used against Gypsy moth (van Netten, 2000 ). These compounds include thietane, acetic acid, 2 - propenyl ester, 2 - butanone, 4 - acetyloxy) - acetic anhydride, trimethylphoshine, benzoic acid, and butylated hydroxy toluene, which are not present in suffi ciently high levels in the gaseous state to pose signifi cant hazard to humans.

15.3.3 Air Fresheners and Deodorizers

Deodorizers are one of the most important groups of consumer products in the modern world. Care products for personal hygiene such as antiperspirants help to avoid the bacterial decomposition of sweat by using preservatives with bacteri-cidal activity. In contrast, odorous VOCs are ingredients of room air fresheners and are released in order to mask unpleasant smells. Solid room deodorizers, which are mainly used in bathrooms and toilet bowl cakes may contain m - and p - dichlorobenzene. Thus, Shields, Fleischer and Weschler (1996) have measured concentrations up to 20 μ g/m 3 of p - dichlorobenzene in commercial buildings. Modern deodorizers are free of halogenated hydrocarbons and are based on iso-prenes such as myrcene, limonene, terpinene and terpinolene (Ohloff, 1990 ). However, most air fresheners for living spaces are sprayed as aerosols. The demand of customers for these products is immense, as is demonstrated by the variety of available fragrances from primrose to violet that abound in commercial stores. Person et al . (1990) identifi ed monoterpenes, ketones, alcohols and alde-hydes as the main components of deodorizers. Measured emission rates ranged from 1.3 × 10 6 − 3.7 × 10 6 μ g/(m 2 h) (toilet deodorizers) to 1.6 10 5 − 2.0 10 6 μ g/(m 2 h) (room fresheners). Shields, Fleischer and Weschler (1996) regarded room freshen-ers and personal care products as the main sources of α - pinene and limonene found indoors. Tichenor and Mason (1988) have monitored nonane, decane, undecane, ethylheptane, limonene and substituted aromatics as representative organic compounds in room fresheners. More recently, Sarwar et al . (2004) showed that the use of liquid and solid air fresheners release ample levels of terpenes indoors and that in the presence of ozone these compounds interact to produce secondary organic aerosols evidenced by the observed increase in particle number and mass concentration in chamber experiments. Therefore, a wide range of consumer products, from air fresheners to cleaners, detergents and perfumes, which contain considerable amounts of limonene, α - pinene and β - pinene, as raw materials have the potential not only to release these compounds into indoor air but to enhance the growth of secondary organic aerosols indoors. Singer et al . (2006) also reported the emission of terpenoids and other volatile organic com-pounds from scented oil air fresheners. For example, the emission rates of the terpenoids ranged from 7.7 mg/day from α - citral to 180 dihydromyrcenol while the corresponding rates for other VOCs ranged from 98 mg/day for 3,7,dimethyl - 3 - octanol to 460 mg/day for benzyl acetate over three days.


Table 15.4 VOC concentrations in a model room after application of 8.1 g room freshener.

Compound Time (h)

0.5 1.0 1.5 2.5 3.3 4.0

Limonene μ g/m 3 29 24 19 18 16 15 Linalool μ g/m 3 38 13 9 6 4 < 1 Nonanal μ g/m 3 13 6 < 1 < 1 < 1 < 1 α - terpineol μ g/m 3 23 7 4 < 1 < 1 < 1 Decanal μ g/m 3 18 5 < 1 < 1 < 1 < 1 Linalylacetate μ g/m 3 96 29 27 14 12 < 1 Ocimene μ g/m 3 13 < 1 < 1 < 1 < 1 < 1 Diethylphthalate μ g/m 3 64 15 7 4 < 1 < 1 TVOC μ g/m 3 502 149 100 72 63 45

In order to determine VOC concentrations after the application of an air fresh-ener (fl owers and fruits), 8.1 g of the air freshener was sprayed into the model room described earlier. The door and windows were kept closed during the test and sampling performed for 0.5 h to 4 h (see Table. 15.4 ). The detected propellents were aliphatic hydrocarbons (C4 and C5). In addition, aliphatic aldehydes (octanal, nonanal, decanal) and terpenes such as myrcene, limonene, linalool, α - terpineol, linalylacetate, ocimene, which are usually present in artifi cial fragrance were detected in trace amounts. The TVOC - value after 0.5 h was 502 μ g/m 3 , which decreased to 45 μ g/m 3 after 4 h. Compared with the background concentration of 33 μ g/m 3 that was measured directly before the test, the air fresheners increased the concentration of VOCs in the room appreciably; the highest concentration of a single compound being 96 μ g/m 3 (linalylacetate).

15.3.4 Cleaning Agents

The number of available cleaning agents for fl oor coverings, furniture, textiles, kitchen, bathroom etc. is enormous. Most of these products are based on inorganic compounds like phosphoric esters, peroxides and surfactants (Raaf, 1992 ). Never-theless, organic dirt such as oil, grease or tar requires special organic cleaning agents. Sack et al . (1992) have investigated more than 100 household cleaners and m - xylene was found in 33% of the samples. In addition, the other main analytes found were acetone, 1,1,1 - trichloroethane, n - octane and methylene chloride. The study of Person et al . (1990) included nine fl oor cleaning products. The main components detected were aliphatic hydrocarbons, aromatic hydrocarbons, oxy-genated compounds and terpenes and the area specifi c emission rates observed ranged from 1.0 × 10 4 − 1.5 × 10 5 μ g/(m 2 h). Colombo et al . (1990) have determined an initial TVOC emission rate of 3.5 × 10 4 μ g/h for a liquid cleaner. A spray cleaner


for carpet emitted 5.0 × 10 4 μ g/h 1 - methoxy - 2 - propanol. Kn ö ppel and Schauenburg (1987, 1989) identifi ed numerous VOCs when investigating a cleanser/detergent. The main components were 2 - methyl - propanol and 2 - propanol and the SER a was 1.1 10 6 μ g/(m 2 h). Sack et al . (1992) found that cleaners for electronic equipment contain mostly chlorinated solvents, 1,1,2 - trichlorotrifl uoroethane, 1,1,1 - trichloro-ethane and methylene chloride being the most prominent. Singer et al . (2006) recently quantifi ed the emissions and concentrations of VOCs in fi ve cleaning agents. When the agents were applied in full strength, the concentrations 1 - h after - use of individual terpenoids measured was found to be in the range 10 – 1300 μ g/m 3 for pine - oil cleaner. However, the corresponding concentration for 2 - butoxyethanol and/or D - limonene for other cleaners was 300 – 6000 μ g/m 3 . In common with terpene - based air fresheners and perfumes, Sarwar et al . (2004) have shown that cleaning agents emit appreciable amounts of terpenoids, which enhance the potential of indoor fi ne particle generation. For example, a general purpose cleaner which emitted a maximum of 20 ppb of terpenes has been shown to increase the particle number in a chamber from 5000 to 211 000 no./cm 3 and the particle mass from 8.1 to 24 μ g/m 3 .

Wolkoff et al . (1998) included, alkanes, halogenated alkanes, alkenes, aromatics, alcohols, glycols, glycol ethers, ethers, aldehydes, ketones, acids and esters among the classes of VOCs typically found in cleaning agents. Of these, terpenoids and glycol ethers were the most frequently encountered in the studies that have been documented. But ingredients of household products are rarely declared on the label. This was especially the case for a paint remover investigated. The only infor-mation given to the customer was that it contained 1 – 2% 1 - methyl - 2 - pyrrolidone. However, this product also contained about 10% of aromatic hydrocarbons (see Figure 15.2 a). Zhu et al . (2005) recently predicted the emission rates of 2 - butoxy-ethanol from 20 consumer products, including cleaning agents. Their results were generally in the range displayed in the Material Safety Data Sheets. This observa-tion is a welcome development that suggests that consumers of this product would be able to understand its potential effects on indoor VOC level and exposure before they purchase it.

As demonstrated by Seifert, Ullrich and Nagel (1989) , emissions from cleaners may result in high VOC levels in the indoor environment. In a private dwelling, up to 450 μ g/m 3 of limonene was measured when a rack used to dry washed linen was kept inside during the winter season. (Limonene is often added to textile softeners used in the washing process.) Clausen, Wilkins and Wolkoff (1998) have investigated house dust for detergents. Fatty acids and fatty acid salts which are common ingredients of fl oor cleaning agents were mainly found.

15.3.5 Polishes

Polishes are used for cleaning, conservation or aesthetic reasons. In most cases, they are applied on large surfaces such as wood, ceramic, marble, linoleum or furniture. Therfore,high emissions of volatile ingredients of polishes from such


Figure 15.2 Volatile ingredients of different household products. A paint remover, B furniture polish, C shoe polish. (1) aliphatic hydrocarbons (C9 - C32); (2) cycloalkanes; (3) branched alkanes; (4) 1 - butanol - 3 - methylacetate; (5) diethyleneglycolmonomethylether; (6) 1 - butanol - 3 - methoxyacetate; (7) C2 - C5 - benzenes; (8) N - methyl - 2 - pyrrolidone; (9) naphthalene; (10) hydronaphthalene.


surfaces may occur readily. Person et al . (1990) have investigated several products and reported that waxes emit large quantities of VOC (1.0 10 6 − 9.4 10 7 μ g/(m 2 h)) when spread on surfaces. Kn ö ppel and Schauenburg (1987, 1989) identifi ed 84 VOCs with SER a s that ranged from 1.2 × 10 5 μ g/(m 2 h) to 2.6 × 10 8 μ g/(m 2 h) in ten different waxes and detergents. Colombo et al . (1990) determined SER u s of 2.7 × 10 4 μ g/h and 1.9 × 10 3 μ g/h for furniture spray polish and fl oor wax paste, respectively while Sparks et al . (1990) have published an emission rate of 1.4 × 10 7 μ g/(m 2 h) for a wood fl oor wax. Because of their high emission rates, waxes for treatment of leather and woollen clothes can cause irritation on inhala-tion, oral or dermal contact.

Organic components of polishes are widespread. Thus most products emit complex VOC mixtures that may consist of alkanes, various alcohols, acetates, C2 - C4 - benzenes, terpenes and derivatives of naphthalene. This is illustrated by the range of compounds emitted by a furniture polish (Figure 15.2 b) and a shoe polish (Figure 15.2 c). Many modern fl oor waxes are based on natural ingredients like alkyd resins. On oxidative degradation of unsaturated fatty acids, volatile aliphatic aldehydes (C5 - C11) with unpleasant smell (Ruth, 1986 ) are formed and the emission rates may remain at high levels over months and even years (Salthammer, 1999 ).

In a recent study, increased use of shoe polish and associated solvents have been found to produce corresponding increases in the levels of benzene, toluene, o - xylene and m - / p - xylene in roadside shoe stalls (Bae, Yang and Chung, 2004 ). Although vehicle emissions and adhesives also contributed to the elevated levels of these compounds, there is little doubt that shoe polishing made signifi cant contributions to the VOC levels. This is evidenced by the statistically signifi cant correlation between shoe polishing and indoor toluene, benzene, o - xylene, and m - / p - xylene at these shoe stalls.

15.3.6 Products for Personal Hygiene and Cosmetics

The exposure to VOCs of humans in the private bathroom is often underestimated. The major sources of exposure to chloroform (CHCl 3 ) in the USA are showers, boiling water and clothes washers. CHCl 3 is formed from the chlorine used to treat water supplies (Ott and Roberts, 1998 ). Similarly, emissions of other by - products of water disinfection, especially trihalomethane s ( THM ), during tap water use and showering constitute important sources of VOC exposure. Several studies (Weisel and Jo, 1996 ; Backer et al ., 2000 ; Nuckols et al ., 2005 ; Chen, Wu and Chang, 2003 ; Xu and Weisel, 2005 ) have suggested that VOC inhalation during showering is a more important contributor to human VOC exposure in the indoor environ-ment than exposure through ingestion. This is probably because of the (i) large amount of water used during showering, (ii) higher temperature of shower water, and (iii) fl ow pattern, which enhance volatilization of VOCs (Nuckols et al ., 2005 ). However, Weisel and Jo (1996) have shown that higher exposure to CHCl 3 occurs by dermal absorption than by inhalation during showering and bathing.


Mathematical modeling of the dynamic behavior of VOCs during showering indicated that the level of VOC exposure risk is determined to a large extent by the type of the showerhead used. Thus jet - fl ow showerheads were found to cause less exposure than spray type showerheads (Chen, Wu and Chang, 2003 ).

Nuckols et al . (2005) investigated the effects of water use activities on blood levels of THMs (such as chloroform, bromodichloromethane, dibromomethane and bromoform) after VOC exposure during hot and cold tap water ingestion, shower-ing, clothes washing, hand washing, bathing and dish washing. Their fi ndings indicated that blood levels of trihalomethanes increased from 57 pg/mL to 358 pg/mL during showering, bathing and hand washing. Therefore, these activities are probably the most serious sources of exposure to water borne VOCs. This is illus-trated in Figure 15.3 by the median THM concentrations measured at a single residence in North Carolina, USA by Nuckols et al . (2005) .

Exposure to VOCs in public beauty shops can also be high. Many cosmetic products contain VOCs such as 2 - phenoxyethanol, 2 - butanone, acetone, terpenes, 2 - hydroxy - 4 - methoxy - benzophenone or phenylmethanol. In particular, hair sprays are potential sources of indoor pollutants. To estimate VOC concentrations associ-ated with the use of beauty products, a female subject was placed in the model room described earlier and sprayed with 16.1 g hair lacquer. Propellant gases (butane, pentane), ethanol, limonene and tripropyleneglycol (isomers) were sub-sequently monitored in the room. Thirty minutes after the application of this product, the highest VOC concentrations were measured for ethanol ( > 100 μ g/m 3 )

Figure 15.3 Median total THM concentration (ug/m 3 ) in exhaled air before and after water activity (Drawn using data from Nuckols et al . (2005) ) .


and limonene (22 μ g/m 3 ). Van der Wal et al . (1997) have studied the exposure of hairdressers to chemical agents. The range of TVOC (C6 - C16) measured was 360 – 660 μ g/m 3 and the main component was always ethanol, which, in the winter months, has stationary indoor concentrations ranging from 4100 μ g/m 3 to 10 400 μ g/m 3 directly after application of sprays. As shown in Table 15.5 , increased concentrations of toluene, limonene and siloxanes were observed in a German hair - dressing shop in April 1998 (door and windows closed). In addition to these compounds, other VOCs such as ethanol, 2 - propanol, n - butanol and various ter-penes were found in the indoor air. Siloxanes are common ingredients of deodor-ants and antiperspirants, and as Shields, Fleischer and Weschler (1996) have pointed out, personal care products are the dominant source of siloxanes, espe-cially decamethylcyclopentasiloxane (D5).

15.3.7 Incenses

Incenses are burned for ceremonial purposes and used as household fragrance. However, emerging facts suggest that they are important sources of many indoor pollutants, including VOCs. One particular study measured emissions from ten types of ‘ traditional ’ , ‘ aromatic ’ and ‘ church ’ incenses in an environmental test chamber (Lee and Wang, 2004 ). The results suggested that burning of incenses, including those claimed to be ‘ environment - friendly ’ may signifi cantly increase indoor concentrations of VOCs in general and concentrations of benzene, toluene, methyl chloride and dichloromethane, in particular as shown in Figure 15.4 . The distribution of the emission factors of carbonyl compounds in the chamber after incense burning can be illustrated by Figure 15.5 .

The suite of VOCs emitted by different types of incenses can vary considerably. For example, methylene chloride increased upon burning traditional incenses, but they decreased when aromatic incenses were burned. Results from another study (Lee, Guo and Kwok, 2002 ) showed that apart from benzene and toluene, considerable amount of 1,3 - butadiene are emitted during incense burning. Thus 1,3 - butadiene increased from below detection limit to about 198 ppbv after burning of incenses (Lee, Guo and Kwok, 2002 ). These studies indicate that incense can

Table 15.5 VOC concentrations in a hair - dressing shop.

Compound Concentration ( μ g/m 3 )

Toluene 288 Limonene 218 Hexamethyldisiloxane 253 Octamethyltrisiloxane 257 Octamethylcyclotetrasiloxane 19


Figure 15.4 Distribution of VOCs emitted by aromatic incenses using data from Lee and Wang (2004) .

Figure 15.5 Mean distribution of carbonyl compounds emitted by six traditional incenses: data from Lee and Wang (2004) .


be a signifi cant source of exposure to toxic organic compounds in the indoor environment.

15.3.8 Perfumes and Fragrances

There is evidence that perfumes have been in use in ancient Egypt since around 2400 BC (Pybus and Sell, 1999 ). More than 4400 years later, their use continues to grow and the list of household and care products which contain perfumes now includes cosmetics, candles, toiletries, air fresheners, laundry products, cleaning agents, foods, beverages, pesticides and fi ne fragrances to name a few (Van Asten, 2002 ). Their importance is also illustrated by the existence of a journal that is dedicated to publishing information that advance knowledge on fl avour and fragrance.

Bridges (2002) suggested that: ‘ The word ‘ perfume ’ substitutes for a detailed listing of fragrance ingredients ’ . Many of these are unsaturated alkenes such as terpenes, ketones such as carvone, esters such as phthalates and musk fragrances such as 1,3,4,6,6,7,8 - hexahydro - 4,6,6,8,8 - hexamethylcyclopenta - (g) 2 - benzopyrane and 7 - acetyl - 1,1,3,4,4,6 - hexamethyl - tetraline (Fromme et al ., 2004 ). But the constitu-ents of perfumes have become so complex and varied that mass spectral analysis is the most useful method of assessing them (Van Asten, 2002 ). However, the use of FID in association with Kovats Index also provides a good insight into the for-mulation of a perfume. Thus a test perfume was shown by FID to consist of several compounds including, benzyl acetate, benzyl benzoate, benzyl salicylate, cerrolide, fl urosa to galaxolide (Van Asten, 2002 ).

The pleasant scents of fragrances profoundly affect the moods and state of well - being of humans. However, a major source of concern is that most ingre-dients of perfumes are VOCs and SVOCs which are readily released into indoor air after the use of the perfume where they can increase water and air pollution and cause or exacerbate adverse health effects. Bridges (2002) reviewed the emerg-ing health and environmental concerns of fragrance and showed that some may be irritants, allergens, sensitizers and phototoxins, suspected hormone disruptors, causative or exacerbating agents for health conditions such as asthma. For example, perfumes were identifi ed in a study as an environmental factor contrib-uting to the development of asthma in children (Bener et al ., 1996 ). In addition, bioaccumulations of fragrances in adipose tissues and breast milk have been reported (Rimkus and Wolf, 1996 ). Inhalation of (+) - limonene and carvone have adverse effects on the blood pressure and many fragrances impact the central nervous system (Bridges, 2002 ). Despite the recognized adverse effects of fra-grances, control and management has been diffi cult because the full composition of fragrances in many consumer products are still ‘ trade secrets ’ known only to the manufacturers, the relevant industry or in some cases government regulatory authorities. For example, fragrance is self - regulated by the industry and in some countries where manufacturers are required to supply the regulatory agencies with information on the composition of consumer products, such


information is not readily revealed to consumers or displayed on labels for con-fi dential reasons.

15.3.9 Cooking and Cooking Related Products

Cooking constitutes a source of VOCs in the indoor environment. Food stuffs and fuels emit ample amounts of VOCs. For example, up to 54 hydrocarbons were identifi ed from a study of 16 fuels/ stoves combinations that are usually used in urban and rural settings in China (Tsai et al ., 2003 ). The worst stove/fuel VOC emitters include metal stoves with a fl ue/unprocessed coal powder, metal stoves with a fl ue/washed coal powder, brick stove with a fl ue/maize residue while the least emitters of VOCs are traditional gas stove/coal gas fuel, improved brick stove with a fl ue/maize residue and metal coal stove with a fl ue/honeycomb coal bri-quette. Many of the compounds emitted in substantial amounts are reactive unsat-urated compounds such as benzene, ethylene, acetylene, and propene. For example, up to 2856 mg of ethylene was emitted by per kg of coal powder in the metal stove. Similarly, the observation of elevated levels of benzene and toluene in a food - court in South China has been rationalized in terms of emission from liquefi ed petrol gas ( LPG ) stoves (Tang et al ., 2005 ). Thus, cooking is an important contributor of precursors of photochemical smog.

In another study (Pandit, Srivastava and Rao, 2001 ), the use of kerosene stoves has increased the concentrations of n - hexane, benzene, heptane, toluene, p - / o - xylene and n - decane in indoors signifi cantly above those in the outdoors. For example, the ratios of indoor/outdoor benzene, heptane and decane ranged from 3.3 to 11.6.

Cooking activity is also associated with increased concentrations of monocyclic aromatic hydrocarbons. Thus, mean concentrations of benzene, toluene, ethylben-zene, m - , p - and o - xylene, styrene, 1,2, 4 - trimethylbenzene and naphthalene were found to be 22.7, 57.0, 6.2, 16.3, 5.6,6.0, 1.4, 5.9 and 0.9 μ g m − 3 respectively in restaurants (Kim, Harrad and Harrison, 2001 ).

15.3.10 Miscellaneous Products and Studies

Wool, which is used as indoor fl oor covers, clothing and in many other consumer products, can emit a wide variety of VOCs, many of which have strong odor. Lisovac and Shooter (2003) have shown that several VOCs can be detected in headspace sampling of wool and wool waxes. A number of these are odorous sulfur - containing compounds while the non - sulfur containing components include hydrocarbons, alcohols, aldehydes and ketones, the most prominent of which are 3 - methylpen-tane, hexane, methylcylopentane, toluene, 2 - methylpentane, ethanol, 1 - butanol, pentanal, hexanal acetone, and 2 - butanone.

Polyurethane , which is an important ingredient in a wide variety of household products including carpet, padding, furniture cushions, sheet foam, varnishes,


sealants, and water sealant products was tested in a chamber and the emission rates of 2, 4 - and 2,6 - toluene diisocyanate ( TDI ) was determined. With the excep-tion of commercially applied water sealant, the TDI emission rates for most of the household products studied was generally < 0.96 μ g/m 2 /h (Kelly, Myers and Holdren, 1999 ). However, for commercially applied concrete sealant, initial emis-sion rates ranged from 257 000 to 360 000 μ g/m 2 /h at 21 ° C and 27 ° C respectively and the 75.2 – 97.8% of the TDIs emitted were the 2,6 - analog.

Similarly, Zhao, Little and Cox (2004) examined PUF as a sink and source of indoor VOCs by investigating its interaction with VOCs such as naphthalene, 1,2,4 - trimethylbenzene, styrene, p - xylene, ethylbenzene, chlorobenzene, toluene and benzene. The results suggest that the sorption of VOCs by PUF is fully revers-ible. Thus it is a potential sink and source of indoor VOCs. (More detailed discus-sion on VOC composition of building products can be found in Chapter 16 .)

Disposal bins have been studied for the VOCs associated with them by Stathero-poulos, Agapiou and Pallis (2005) . The most prominent classes of compounds emitted are generally aliphatic and aromatic hydrocarbons, esters, terpenes and alcohols. The highest median concentrations for a single compound was 649.9 μ g/m − 3 for decane while median concentrations of several other compounds such as limonene, undecane, nonane, ethanol, acetic acid ethyl ester and 1, 2,4 - trimethylbenzene are in the range 159.1 – 353.1 μ g/m − 3 . Therefore, waste bins are not only sources of odorants but sources of VOC as well.

Personal computer s ( PC s) are important sources of VOCs in offi ce and homes (Bako - Biro et al ., 2004 ).Thus, the TVOCs emission rate per PC observed in a glass chamber study was as high as 486.6 μ g/h while individual emission rates for toluene and phenol were 47 and 63 μ g/h respectively. Other prominent chemicals emitted by PCs include, 2 - ethylhexanol, formaldehyde and styrene (Bako - Biro et al ., 2004 ). (See Chapter 17 for a more detailed discussion of VOCs in electronic devices.)

Odor analysis (see Chapter 8 ) performed using GC coupled with olfactometry has also shown that many food items and household materials are odorant sources (Mayer and Breuer, 2006 ). Thus, mono - unsaturated aldehydes particularly E - 2 - nonenal are found in fat, wax, oil fi nish and lubricants; branched aldehydes such as 3 - methyl butanal are found in varnish, bread and malt; while leather, rice and popcorn are sources of substituted pyrrolines especially 2 - acetyl - 1 - pyrroline. Studies like this are important not only from the point of view of identifying sources of indoor odorants but also from the point view of providing vital informa-tion that can help consumers to select products.

Tang et al . (2005) have reported signifi cant levels of VOCs in many supermarkets and department stores. In particular, the ratio of indoor to outdoor concentrations of 1 - 4 dichlorobenzene was as high as 39 in a store where leather products are used, 77.8 in a store that sells luxury fashion items and clothing, and 39 in a supermarket that sells household supplies. These observations demonstrate that consumer goods are potential sources of indoor VOC exposure.

Studies of the indoor air quality in aircraft cabins have indicated that their VOC levels are similar to those of residential houses. The only notable differences


concern the levels of ethanol, acetone, and chlorinated hydrocarbons. While the ethanol and acetone levels in aircraft cabins are higher than those in residential/offi ce buildings, the levels of chlorinated hydrocarbons such as dichloromethane in residential and offi ce buildings are higher than those in aircraft cabins (Nagda and Rector, 2003 ). The presence of such appreciable levels of the former set of compounds in aircraft possibly refl ects their emissions from biological effl uents and consumer products. In addition to consumer products, ‘ disinsection ’ signifi -cantly enhances the VOC levels in aircraft cabins (Berger - Prei β et al ., 2006 ). For example, 5 mins after spraying the cabin with ‘ disinsectant ’ , the concentrations of D - phenothrin varied from 853 – 1753 μ g/m 3 for an Airbus 311 and a Boeing 747 - 400. The concentration decreased to 36 – 205 μ g/m 3 5 – 20 mins later and 1 μ g/m 3 20 – 40 min later.

15.4 Conclusion

Volatile organic compounds are emitted from a wide variety of household and consumer products. The emission rates are distributed over several orders of magnitude and are strongly dependent on the mode of application of the products. In many cases, VOC mixtures are emitted but a suite of organic components can usually be related to individual product classes (Tichenor and Mason, 1988 ). Chamber experiments and simple headspace analysis are particularly useful in this regard (Colombo et al ., 1990 ). Nevertheless, it is diffi cult to identify sources of indoor air pollution and to estimate human exposure, because volatile ingredi-ents are not always declared on the labels of goods. In the modern world, where several products are available to the customer to choose from, necessity and mode of application need to be critically considered when goods are selected for indoor use. Ott and Roberts (1998) stated: ‘ Yet people cannot take the simple steps required without adequate knowledge. So increased education is needed. Law requiring more detailed information would also help: If a product contains a dangerous pollutant, should not the manufacturer be required at least to list the chemical by name on the package? Armed with a better understanding of toxic substances found in common products and in other sources at home, people could then make their own informed choice. ’

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Building Products as Sources of Indoor Organic Pollutants Stephen K. Brown

16

16.1 Introduction

This chapter will consider the role of building products as major sources of organic indoor air pollutants. The organic pollutants will include VOCs and formaldehyde, the latter since formaldehyde is a prevalent indoor air pollutant from many building products. While it is well - established that building products are the most signifi cant sources for VOCs and formaldehyde in new buildings (Brown et al ., 1994 ; Brown, 1999a ), leading to signifi cant levels of pollution (Marbury and Kreiger, 1991 ; Dingle, Murray and Jiang, 1992 ; Brown, 1997 ), less is understood about long - term, low - level emissions which are often assumed to be insignifi cant. A life cycle assessment approach to this subject is relevant since the impacts on building occupants occur over a building ’ s life, and so the possibility of longer term emissions should also be considered. This chapter will consider short - and long - term emissions of primary organic pollutants, that is, those emitted from the material itself in the absence of other pollutants or environmental factors. Current research on oxidant reactions with indoor materials leading to ‘ secondary ’ organic emissions is also a life cycle approach, as is material degradation by environmental factors (heat, moisture, light). However these are beyond the scope of this chapter, are discussed elsewhere, or will receive limited attention only when relevant to the general discussion.

16.2 Organic Pollutants Emitted from Major Building Products

16.2.1 Building Products

Building products are considered to be those products used internally within a building during its construction and which are fi xed in place as part of the construc-


373

374 16 Building Products as Sources of Indoor Organic Pollutants

tion process. Even with this distinction, there may be different perceptions of what constitutes a building product. For example, a fl oor is an integral part of a building but when it is covered, is it considered an internal product or not? Is a fl oor cover-ing a building product, since it is fi xed in place but will be periodically replaced? This review uses ‘ building product ’ in a broad sense, and it is more useful to state what is not considered to be a building product – materials within wall or fl oor cavities, moveable furniture and furnishings, offi ce equipment, and mobile appli-ances are not considered to be building products. In addition, many of the building products considered here are large - area materials in buildings since it is logical to expect these to have greater potential to be signifi cant pollutant sources. Lastly, mostly organic materials are discussed; inorganic materials (glass, metal, stone, cement/concrete, fi red tiles) are nonemitting for organic compounds unless coated or treated with an organic material.

16.2.2 Organic Pollutants

Volatile organic compounds were defi ned by the World Health Organization (WHO, 1989 ) as all organic compounds occurring in the boiling point range of 50 – 100 ° C and 240 – 260 ° C. The temperature ranges at these extremes were selected since adsorbents were variable in capturing or releasing some VOCs, but with current technology it is common to express one range, 50 – 260 ° C or specifi c VOCs according to their GC elution at these extremes. Organic compounds with lower boiling points were considered very volatile (VVOC) and those with higher boiling points as slightly volatile (SVOC). VOC alone encompasses a very large number and diverse range of organic compounds, which will be referred to using names from the International Union of Pure and Applied Chemist s ( IUPAC ) nomencla-ture. Formaldehyde, which has a boiling point of − 20 ° C, is defi ned as a VVOC. It is included in this chapter because it is considered to be an important organic pollutant in building product emissions to indoor air environments (Brown, 1997 ).

While it is clearly important to consider exposure to individual toxic pollutants, it has been postulated that exposure to a mixture of VOCs, individually at low concentrations but at much higher concentration as an aggregated ‘ total VOC ’ (TVOC), can cause sensory irritation and discomfort (M ø lhave, 1990 ; M ø lhave and Nielsen, 1992 ). TVOC concentration is essentially the summed concentration, in units of μ g m − 3 , of all of the VOCs present in the air sample, although many studies have used different defi nitions (e.g., specifi c VOCs, different air sampling methods, different measurement instruments, different quantitation units), which have large impacts on the TVOC concentration value (Hodgson, 1995 ). A specifi c TVOC defi nition has been developed by the European Commission (1997) , based on a list of 64 VOCs commonly found in indoor air (Table 16.1 ) plus any other VOCs present in large quantities. The VOCs emitted from materials in Australia have shown a high level of consistency with this list (Cheng and Brown, 2005 ). Note that when referred to in this chapter, VOCs from this list (and formaldehyde) will appear in bold . However, analytical procedures to measure TVOC to this defi nition

16.2 Organic Pollutants Emitted from Major Building Products 375

Table 16.1 64 VOC s of importance to indoor air and included in TVOC ( EC , 1997 ).

VOC VOC VOC

Benzene n - pentadecane Butanal Toluene n - hexadecane Hexanal Ethylbenzene 2 - methylpentane Nonanal m - xylene 3 - methylpentane Pentanal p - xylene 1 - octene Benzaldehyde o - xylene 1 - decene Ethylmethylketone (EMK) n - propylbenzene Methylcyclopentane 4 - methyl - 2 - pentanone 1,2,4 - trimethylbenzene Cyclohexane Cyclohexanone 1,3,5 - trimethylbenzene 3 - carene Acetophenone 2 - ethyltoluene α - pinene Trichloroethylene Styrene β - pinene Tetrachloroethylene Naphthalene Limonene 1,1,1 - trichlorethane 4 - phenylcyclohexene 2 - propanol 1,4 - dichlorobenzene n - hexane 1 - butanol Hexanoic acid n - heptane 2 - ethyl - 1 - hexanol Ethyl acetate n - octane Ethylene glycol methyl ether n - butyl acetat n - nonane Ethylene glycol ethyl ether Isopropyl acetate n - decane Ethylene glycol butyl ether 2 - ethoxyethyl acetate n - undecane 1,2 - propylene glycol methyl ether 2,2,4 - trimethylpentane - diol diisobutyrate (TXIB) n - dodecane Diethylene glycol butyl ether tetrahydrofuran n - tridecane 2 - pentyl furan n - tetradecane Methylcyclohexane

are limited (Massold et al ., 2001 ). This chapter will refer to some studies of TVOC emissions from building products, but without discriminating the differences likely to be present due to TVOC defi nitions.

Inclusion of SVOCs and VVOCs is beyond the scope of the chapter, but some attention to these will be paid where they constitute signifi cant emissions from building products.

Only pollutant emissions determined under well - controlled conditions (e.g., in environmental chambers) will be presented since the critical aim of this chapter is the understanding of volatile organic pollutant emissions from specifi c building materials sources.

16.2.3 VOC Emissions Levels Over Time

VOC emissions from building materials can vary over time by a range of behavior, generally related to the underlying mechanisms of VOC transport in the materials. Some emission behavior can be described as:


• decreasing from high initial levels to near - zero levels (e.g., fi rst - order exponen-tial decay of some wet paints and coatings);

• decreasing from high initial levels to near - constant (or slow decay) levels (e.g., double - exponential decay of VOCs from some carpets and fl oor coverings; form-aldehyde emission from reconstituted wood - based products [Brown, 1999b ]);

• increasing from near - zero levels to pass through a peak emission several hours after application/manufacture (e.g., delayed fi rst - order emission as in aldehyde formation by auto - oxidation of fatty acids in alkyd paints, vegetable oil products [Chang et al ., 1999 ]);

• occurring at a constant emission rate until the source has depleted or is inactivated (e.g., equipment operation (Brown, Mahoney and Cheng, 2004 ).

Note that this behavior relates to single VOCs. Since organic compounds will have different physico - chemical properties (especially volatility, polarity, molar volume, hydrolytic stability) that infl uence their emission behavior, it is quite possible for a material to emit one group of dominant VOCs in its early emissions and a dif-ferent group later (e.g., when a wet paint fi lm has ‘ dried ’ ), in each case at somewhat different orders of emission rates.

This chapter will not delve into emission mechanisms. However, building product emissions should be considered in relation to building occupancy/life cycle factors by:

• early emissions (0 – 12 hours after application or manufacture); • mid - term emissions (1 – 14 days); • late emissions (two or more weeks); • long - term emissions (three or more months).

16.2.4 VOC Emission Limits/Labels

Many emission limits exist for building products in different countries. By their nature, these limits aim to link VOCs that are detrimental to indoor air quality to specifi c materials that act as their sources in buildings. These VOCs will be considered in the discussion of each building product.

Also, many research papers and data sources express these emissions differently as:

• chamber concentration after time t ( C t μ g m − 3 ) under the conditions of test; • an emission factor after time t ( EF t μ g m − 2 h − 1 ; pollutant mass/area of surface/

time) for the source product.

Wherever possible, this chapter will present data as EF since this allows a com-parison of product emissions, but these can be converted to C using the product loading ratio ( L m 2 m − 3 ) and ventilation rate ( N h − 1 ), as follows:

C L Nt tEF= ⋅ (16.1)

16.3 Interior Paints 377

16.2.5 TVOC Emissions from Building Materials

In a previous review, there were clear differences in TVOC emissions from build-ing products (Brown et al ., 1994 ), with ‘ wet ’ materials (paints, adhesives, sealants) of greatest signifi cance, occurring at orders of magnitude higher emissions than other building products. The review also showed that indoor TVOC levels were often an order of magnitude greater in new buildings ( < 3 months old) than in established buildings, consistent with the high impact on IAQ of emissions from new building products.

TVOC emissions can allow a broad comparison of the indoor impacts of build-ing products. TVOC is used in many product emission criteria, the most recent of which (AgBB, 2002 ) sets a 3 - day criterion of 10 m g m − 3 and 28 - day criterion of 1 m g m − 3 , as chamber concentrations, which are equivalent to emission factors of 10 and 1 m g m − 2 h − 1 , respectively. Table 16.2 presents data from CSIRO ’ s laboratory which should allow comparison of relative differences of building products and a limited comparison with the 3 - day criterion (note that 28 - day emission measure-ments were not available). Solvent - based paints were clearly the highest emitters and could exceed the 3 - day criterion. Water - based acrylic paints show signifi cant variability in emissions, and very low emissions when produced as low - VOC products. Natural paints were higher emitting than the low - VOC paints. Carpet/underlay, fl oor coverings, and wood - based panels all exhibited order of magnitude variations within each product range, but were broadly similar in TVOC emissions to the low - VOC paints and much below the AgBB 3 - day criterion.

Generally, the conclusions of the previous review are supported on the basis of TVOC emissions. However, the nature, levels and persistence of the specifi c organic emissions from each product range are the key factors to consider for IAQ impact, and are the focus of this chapter.

16.3 Interior Paints

Interior paints consist of a complex range of products but can largely be considered as water - based and solvent - based paints which have been used for decades, and low - emission and ‘ natural ’ paints that have been developed over the last decade. Up to the mid - 1990s there was a trend to replace solvent - with water - based paints, which Norback, Weislander and Edling (1995) reported had reduced the exposure levels of painters to TVOC by a factor of 100. However, they also noted this shift resulted in exposures to VOCs that were more polar, with higher boiling points and with poorer understanding of health impacts but potential for mucous irrita-tion and airway infl ammation. The USEPA (Chang, 2001 ) summarized knowledge on paint hazards and indoor air as follows:


Table 16.2 TVOC emission factors from building products assessed in CSIRO ’ s laboratory.

Building products TVOC emission factor ( m g m − 2 h − 1 )

1 day 3 days

Solvent - based paints Polyurethane (1 day old) 20 12 Alkyd enamel 130 8

Water - based paints Acrylic latex 1 80 8 Acrylic latex 2 5.8 1.5 Acrylic latex 3 (low odor) 5.0 0.4

Low - VOC paints Zero - VOC acrylic 1 0.6 < 0.04 Zero - VOC acrylic 2 0.2 0.1 Zero - VOC acrylic 3 0.01 < 0.01 Low VOC acrylic 4 1.1 0.7

‘ Natural ’ paints Vegetable oil paint 0.2 0.8 Orange peel extract paint 0.4 0.08

Carpet/underlay Wool carpet 1 1.9 1.4 Wool carpet 2 0.9 0.4 Wool carpets (3 – 5) 0.1 – 0.2 – Wool carpets (6 – 8) < 0.1 – Polyurethane foams (4) 0.05 – 0.2 –

Floor coverings Linoleum – 0.1 Polyolefi n – 0.1 UV - cured lacquer/timber 0.4 0.3 Rubber – 0.6

Wood - based panels MDF 1 < 0.04 – MDF 2 1.2 – Zero - formaldehyde MDF 0.3 – Particleboard 0.8 –

Heating/cooking Ceiling heat panel 0.5 a – Unfl ued gas heaters (4) 0.02 – 0.06 a Electric ovens (3) 0.2 – 1.6 a

a Emissions in m g m − 3 h − 1 per appliance in operation.


• For solvent - based alkyd enamel paints: � Based on emission chamber work, the approximate 8 - h time - weighted average

(TWA) for TVOC levels can exceed 2000 m g m − 3 . � Xylene emissions from some paints approached the exposure levels at which

neurological effects are often seen. � These high solvent exposures lead to concern about the possibility of chronic

central nervous system effects in professional painters. � TVOC exposure levels indicate a high likelihood of complaints about indoor

air quality during and shortly after painting. � For paints emitting 2 - butanonoxime, the risk posed by the maximum

occupational exposure during typical use presents concerns regarding developmental toxicity health effects.

� The risk of cancer among consumers and professional painters from the inhalation of 2 - butanonoxime during paint application and drying was con-sidered to be a concern.

• For water - based latex paints, acute and chronic risks exist for professional painters based on exposure to emissions: � TVOC levels were in the range that may result in complaints about indoor air

quality. All samples and test conditions in the chamber studies resulted in TVOC levels exceeding 40 mg m − 3 as a 24 - h average.

� Some latex paints were sources of formaldehyde and acetaldehyde. � Acetaldehyde exposure presented a chronic health risk and a concern for

cancer risk with (lifetime) risk estimates of 10 − 4 based on potency estimates.

� Formaldehyde exposure raises a marginal concern for cancer risk to profes-sional painters (risk estimates of 10 − 4 to 10 − 6 ).

� Formaldehyde also presented an acute irritation concern for both consumers and professional painters since its emission could persist for more than one month.

16.3.1 Water - Based Paints

Water is a poor solvent for organic resins and so water - based paints are generally water - reducible resins (where the resin is dissolved in an alcohol and then diluted with water) or more commonly polymer latexes:

• Water - reducible resin coatings consist of an acidic or basic polymer, an auxiliary solvent, a neutralizing agent and water. Auxiliary solvents include 1 - butanol , 2 - butanol, ethylene glycol butyl ether and propylene glycol ethers . Ammonia, amines and amine derivatives are used as neutralizing agents (e.g., triethyl-amine, dimethylethanolamine, dimethylaminomethylpropanol) (Doren, Freitag and Stoye, 1994 ).


• Water - based latex paints are more commonly used than water - reducible resins for interior paints, and contain a range of organic solvents to ensure wet fi lm stability during drying, coalescence of latex particles during late stages of drying to ensure fi lm formation, and escape of solvent residues after coalescence to reduce fi lm softness. The remaining discussion is focused on these paints.

Typical VOC contents in water - based latex paints are approximately 180 g l − 1 . These consist of cosolvents to regulate paint drying rates, and coalescing solvents within the latex particles to ensure the particles merge to yield a cohesive fi lm. Cosolvents include ethyl methyl ketone , 2 - propanol , propoxypropanol, ethylene glycol butyl ether and diethylene glycol ethyl ether. Coalescing solvents are usually glycol ethers and esters, which reduce the glass transition temperature of the polymer particles ( T g , the temperature above which a long - chain polymer molecule or seg-ments of the molecule attain molecular mobility) to below ambient temperature. However, this means the polymer fi lm will be soft and retain dirt. The coalescing solvent must evaporate after fi lm formation, resulting in a paint fi lm with a T g well above ambient temperature (Wicks, Jones and Pappas, 1992a ).

Other volatile constituents of interior water - based latex paints are:

• in - can biocides, including formaldehyde ;

• supplemental fungicides such as aromatic mercury compounds (e.g., phenyl-mercuriacetate which causes elemental mercury emission (Tichenor, Guo and Dorsey, 1991 ) although these compounds are prohibited in some countries; and

• ethylene glycol and 1,2 - propylene glycol, as ‘ lapping ’ (the ability to brush paint into a partially dried fi lm without damaging it) and antifreeze agents.

VOCs found in water - based paints in the Netherlands and Denmark (by a survey of manufacturers) are presented in Table 16.3 . VOC emission assessments have shown that most of these VOCs are emitted from water - based paints, though with signifi cant variation from product to product in type and quantity of VOC emitted. For example, Table 16.4 presents the VOC emissions from wet products under identical test conditions (Brown, 2000 ). WBP1 has no 1,2 - propylene glycol while the other paints all use large quantities, apparently replacing the glycol ethers and esters.

Building assessments have identifi ed these VOCs in established buildings, sug-gesting their long - term release from painted surfaces (though other uses, such as in cleaning materials, were also possible):

• Bluyssen, Fernandes and Groes (1996) identifi ed propylene glycol ethyl ether, ethylene glycol butyl ether , ethylene glycol phenyl ether, diethylene glycol ethyl ether, acetic acid butyl (or ethyl) ester , butoxyethoxyethyl acetate in established European offi ces.

• Pleininger and Marchl (1999) identifi ed ethylene glycol ethyl ether, ethylene glycol butyl ether , ethylene glycol phenyl ether, 2 - n - butoxyethylacetate, 2 - (2 - methoxyethoxy)ethanol , diethylene glycol ethyl ether, diethylene glycol


Table 16.3 VOC s in water - based paints from the Netherlands and Denmark (Hansen, Larsen and Cohr, 1987 ; Van Faassen and Borm, 1991 ) .

Compound CAS No. Max. Content %w/w

Netherland Denmark

White spirit ( ≥ 80%w/w C 9 - C 11 hydrocarbons; aliphatics, alicyclics, aromatics)

64742 - 88 - 7 3.6 2.9

Esters and others Isobutyl esters of dicarboxylic acids – 2.0 1.0 2,2,4 - Trimethyl - 1,3 - pentanediol

monoisobutyrate (Texanol ® ) 25265 - 77 - 4 5.0 5.0

N - Methyl - 2 - pyrrolidine 872 - 50 - 4 1.7 –

Glycols Ethylene glycol 107 - 21 - 1 7.9 2.0 1,2 - Propanediol 57 - 55 - 6 7.9 10

Glycol ethers Ethylene glycol ethyl ether 110 - 80 - 5 2.0 1.1 Ethylene glycol phenyl ether 122 - 99 - 6 2.5 5.0 Ethylene glycol butyl ether 111 - 76 - 2 3.0 1.4 Diethylene glycol ethyl ether 111 - 90 - 0 3.0 1.0 Diethylene glycol butyl ether 112 - 34 - 5 5.0 1.5 Propylene glycol - 1 - methyl ether 107 - 98 - 2 2.1 – Dipropylene glycol - 1 - methyl ether 34590 - 94 - 8 7.0 4.0 Butylene glycol - 3 - methyl ether 2517 - 13 - 3 5.0 –

butyl ether , 2 - (2 - n - butoxyethoxy)ethylacetate, 1,2 - propylene glycol, 1,2 - propylene glycol methyl ether , propylene glycol butyl ether, propylene glycol phenyl ether, 2 - ethyl - 1 - hexanol , 1 - methyl - 2 - pyrollidone, Texanol in established residences in Germany.

IAQ emission criteria for water - based paints have been suggested as follows:

• Larsen and Abildgaard (1995) : at 1 – 3 hours after painting VOC emis-sion < 3500 μ g m − 2 h − 1 ; at 4 weeks after painting VOC emission < 7 μ g m − 2 h − 1 ; VOCs to include glycols, glycol ethers, Texanol, and TXIB . No recognized car-cinogens to be emitted.

Much of the research into VOC emissions from paints has been carried out using impermeable substrates (glass, metal) rather than the porous substrates used in buildings. This can have a signifi cant infl uence on the quantity and duration of VOC emissions. Tichenor (1995) described the research with a polyvinylacetate latex paint applied to stainless steel or gypsumboard and evaluated for 7 days.


Table 16.4 VOC emissions from four conventional water - based paints (23 ° C, 50% RH , 1.0 h − 1 , 0.5 m 2 m − 3 ).

Paint VOC EF ( μ g m − 2 h − 1 ) at

8 h 1 d 3 d 14 d

WBP1 m,p - Xylene 42 < 12 < 2 < 1 1,2 - propylene glycol ethyl ether 72 000 44 000 4000 < 2 n - Butylether 300 36 < 2 < 2 Ethylene glycol butyl ether 380 < 60 < 4 < 2 Texanol a 38 000 13 200 3000 220

WBP2 Acetone/ n - pentane 1 700 60 2 < 2 Isobutanal 780 < 24 68 10 1,2 - Propylene glycol 110 000 9 000 < 12 < 8 Texanol ® a 10 000 3 600 1900 160

WBP3 Acetone 1 900 300 86 24 1,2 - Propylene glycol 82 000 5 000 < 2 < 2 Diethylene glycol butyl ether 8 000 2 600 340 8

WBP4 Isobutyraldehyde 640 – 64 < 2 1,2 - Propylene glycol 62 000 3 000 140 10 Texanol ® a 16 000 10 000 2000 40

a Texanol ® = 2,2,4 - trimethyl - 1,3 - pentanediol - isobutyrate.

Emissions factors from painted gypsumboard were considerably lower due to its sorbency properties In a long term (6 month) experiment on the painted gypsum-board, most VOCs (1,2 - propylene glycol, diethylene glycol butyl ether , Texanol ® ) decayed to close to their limits of detection after 2 – 3 months, while ethylene glycol persisted at levels 15 – 20 times higher. It was found that a gas - phase mass transfer model could be fi tted to data to 7 days but not to long - term measurements, which followed a second order model (rate constant k 2 ):

EF EF EFt o o= ( ) + ( )( )1 2k t (16.2)

This was considered to indicate that long - term emissions were controlled by VOC diffusion processes that occurred in the gypsumboard.

Wilkes, Koontz and Cinalli (1996) investigated the emission during 8 to 9 days of low vapor pressure VOCs from water - based paints applied to prepainted gypsumboard. They observed that a double exponential decay model (empirical constants a , b , x , y ) fi tted the data well:

c a e x e= −( ) − −( )( )− −1 1bt yt (16.3)


A fast initial emission decay up to one day was correlated to vapor pressure, indi-cating an infl uence of evaporative processes. A slow emission decay from one to nine days was correlated to molecular weight indicating control by fi lm diffusion processes. However, they found that the fast initial release generally accounted for less than 10% of the total released VOCs. Also, they found that only 20 to 35% of the applied VOC mass was released, that is, there was long - term retention of VOCs by the gypsumboard.

Water - based paints may also contain formaldehyde as an in - can biocide. This will also be released as the paint dries, though this appears to have received limited research. Brown (2000) measured formaldehyde EF of < 20 – 300 μ g m − 2 h − 1 from 3 – 24 hours after applying the water - based paints in Table 16.4 , the higher EF being observed at earlier times. Colon and Mookherjea (1997) reported formaldehyde emissions from water - based paints using two formaldehyde - based biocides and applied to gypsumboard. They predicted that in a typical room ventilated at 0.5 h − 1 the formaldehyde concentration would peak after 1 – 3 hours at 40 – 70 μ g m − 3 and by 24 hours would be less than 10 μ g m − 3 . The observations with low - VOC paints (see later) show these predictions may underestimate the persistence of formalde-hyde after painting.

16.3.2 Solvent - Based Coatings

Solvent - based coatings are typically solution systems such as alkyd enamel paints and polyurethane lacquers. Blends of up to ten solvents can be used, which are selected to ensure resins dissolve and stay dissolved during drying, as well as to control the rate of drying. A VOC content above 275 g l − 1 is generally used. The organic solvents used in the coatings can be classifi ed into three broad categories (Wicks, Jones and Pappas, 1992b ):

i) Weak hydrogen - bonding: • aliphatic and alicyclic hydrocarbons – examples are naphthas which provide

high volatility for lacquers, and mineral spirits for slower evaporating coatings;

• aromatic hydrocarbons, for example, toluene and xylene , which dissolve a broader range of resins and are used on a large scale; and

• chlorinated hydrocarbons (the use of which is increasingly limited due to toxicity concerns and ozone depletion).

ii) Hydrogen bond acceptors (ketones and esters): • ethyl acetate , isopropylacetate , n - butyl acetate , ethyl methyl ketone , isobutyl

methyl ketone and methyl - n - amyl ketone are those most widely used. Slow - evaporating esters are 1 - (methoxy - 2 - propyl)acetate and 2 - butoxyethylacetate.

iii) Hydrogen bond donor - acceptors (alcohols): • methanol, ethanol, 2 - propanol , 1 - butanol , 2 - butanol and isobutanol are

those most widely used.


Solvents are selected such that some will escape relatively quickly from paint fi lms to prevent excessive fl ow, while others will escape slowly to provide fi lm leveling and adhesion. With typical alkyd coatings, the fi rst 30% of solvent has been observed to evaporate as quickly as the neat solvents, essentially at a constant rate which is dependent on volatility. Later stage evaporation occurred several times more slowly and was rate - controlled by solvent diffusion to the surface of the paint fi lm. The transition point between this behavior was defi ned as the resin solids content at which the evaporation rate due to volatility equaled that due to diffusion. Transition points have been observed to typically occur at a resin solids content of 40 – 50% v/v. Thus alkyd paints, normally formulated at 27 – 40% v/v resin solids, generally exhibit rapid initial solvent - release driven by volatility; while high solids coatings (usually 65 – 75% v/v resin solids) dry solely by a diffusion - controlled process with negligible infl uence by solvent volatility (Ellis, 1983 ).

Solvent loss in the diffusion - limited phase is expected to be important to the medium - and long - term solvent emission properties of coatings. In the medium term, solvent diffusion in the (still) liquid coating will become slower than surface evaporation as viscosity increases with solvent loss. In the long term, solvent dif-fusion in a nonliquid fi lm may continue for very long periods, depending on molecular properties of the resin and solvent. Hansen (1968) observed that for late - stage (low solvent concentration) drying, the diffusion coeffi cients of solvent molecules in polymers were exponentially related to solvent concentration in the polymer. He found that solvent evaporation from polyvinylacetate during the fi lm diffusion phase was described by a reduced time variable D o t � − 2 for several sol-vents, where D o is the diffusion coeffi cient at zero solvent concentration, t is time and � is fi lm thickness. He concluded that hydrogen and polar bonding in the fi lm were not important to the process, with the critical factor being the solvent diffu-sion coeffi cient. Newman and Nunn (1975) investigated solvent retention and diffusion processes in organic coatings and noted that some solvents were retained in polymer fi lms at high levels (e.g., 10 – 20% solvent) for periods much in excess of several months. No direct relation existed between the extent of retention and solvent volatility in single solvent systems. They found that the order for solvent retention was largely independent of the type of polymer and the solvent volatility. Instead, solvents of low molecular volume (the volume occupied by one mole of the compound, cm 3 /mole) tended to be least retained. Newman and Nunn (1975) considered these effects were explained by a rate - determining mechanism of dif-fusion of solvent molecules through the polymer fi lm.

As with water - based latex paints, the substrate on which the solvent - based paint is applied will infl uence the emission rates of VOCs. Kwok et al . (2003) showed that emission rates of aromatic VOCs from a polyurethane lacquer in the fi rst 10 hours were 65% greater when the substrate was aluminum rather than plywood. VOCs from the former were emitted in an evaporation process, while from the latter they emitted largely in a diffusion - controlled process.

An additional complexity for alkyd enamel paints is mid - term emissions. Table 16.5 shows the emissions from a typical full - gloss alkyd enamel paint (Brown, 2000 ). Note the very high early emissions of the solvents (especially c.f. the


Table 16.5 VOC emissions from a solvent - based paint: alkyd enamel full - gloss (23 ° C, 50% RH , 1.0 h − 1 , 0.5 m 2 m − 3 ).

VOC Emission factor ( μ g m − 2 h − 1 ) at

8 h 1 d 3 d 14 d

Toluene < 100 16 4 < 2 Ethylbenzene 140 < 18 < 8 < 2 o,m,p - xylene 2 600 < 30 4 < 2 3 - Methylethylbenzene 4 600 < 60 < 8 < 4 1,2,4 - Trimethylbenzene 13 000 46 10 2 n - decane 24 000 38 8 < 2 n - undecane 110 000 460 12 < 2 n - dodecane 54 000 8 400 60 < 2 Hexanal 680 11 000 1 600 120 Heptanal < 600 260 80 14 Octanal < 3000 84 40 32 TVOC 840 000 76 000 13 000 1400

water - based paint emissions), followed by the increasing emission of the aldehydes hexanal , heptanal and nonanal . Most of the VOC emissions from this paint occurred in the fi rst four hours as the paint became ‘ dry ’ , when the aromatics and alkanes were emitted. Peak concentrations for each of the aromatic and alkane VOCs generally occurred at longer times as VOC volatility decreased showing a dominant role of compound volatility in early emissions. Emissions of hexanal , heptanal and octanal before 4 to 8 hours were diffi cult to quantitate due to the much greater quantities of other VOCs present in the air samples. However, it appeared that the aldehydes started to be emitted after 8 hours, when the paint was dry and most of the aromatic and alkane VOCs had been emitted, even though they were much less volatile than the aldehydes. The aldehyde emissions were considered to have resulted from autoxidation reactions of the unsaturated fatty acids in the alkyd paint (Wicks, Jones and Pappas, 1992a ) after most of the solvents had been emitted and an air interface had been established with the alkyd resins. The aldehyde emissions then occurred over a period of several days as the alkyd resin cross - linked to a hard coating. As a result of this process, hexanal was the major VOC emitted from the coating after one day, by which time the emissions of aromatics and alkanes had virtually ceased. Hexanal is the main aldehyde emitted in the auto - oxidation reaction of methyl linoleate, the fatty acid in greatest abundance in linseed oil, a common component of alkyd paints. Chang and Guo (1998) determined that hexanal was emitted from alkyd paints by an autoxidation reaction that could be described by the following consecutive fi rst order model:

Alkyd Intermediate Hexanal2k k1⎯ →⎯ ⎯ →⎯


where k 1 and k 2 are fi rst order rate constants. Assuming there was a delay period t a before initiation of the reaction, they found that hexanal emissions were described by the model:

EF EFt a at a at= − −( )( ) − − −( )( )[ ]0 1 2exp expk t t U k t t U (16.4)

where

EF 0 = k 1 k 2 A 0 /( k 2 – k 1 ) U at = unit step function that equals 1 if t > t a , otherwise = 0.

They observed that values of k 1 were much smaller than values of k 2 , showing that hexanal emission was predominantly controlled by the reaction forming the intermediates.

16.3.3 ‘ Natural ’ Paints

A number of interior paints that use ‘ natural ’ ingredients, as opposed to synthetic chemicals, are commercially available, though little has been published on the physical properties and pollutant emissions of such products. They are believed to be used by chemically sensitized people rather than the general population, but no study is known of the health signifi cance to the former from such products. The Canada Mortgage and Housing Corporation (1995) summarized these prod-ucts as follows:

i) Natural oil - based paint – most contain no petroleum products but some may contain small amounts of de - aromaticized petroleum solvents (isoparaffi ns); citrus oil and linseed oil odors may be present during drying and may cause irritation; some natural ingredients can cause problems for certain allergic individuals; personal evaluation to determine if ill - effects are associated with indi-vidual paint is recommended .

ii) Natural water - based paint – casein - based paints that may release ammonia during application; as for (a), because of potential individual sensitivity, per-sonal evaluation of ill - effects is recommended .

Brown (2000) evaluated VOC emissions from two of these products, as summa-rized in Table 16.6 . Paint 1 was claimed to use ‘ orange peel oil ’ as its base. It exhibited very high and fast - decaying emissions of C 7 - C 10 alkane and limonene , with EF at 2 hours of 70 000 and 120 000 μ g m − 2 h − 1 , respectively. Paint 2 was claimed to be based on ‘ vegetable oils ’ and was virtually nonemitting at application, but emitted several malodorous aldehydes (including formaldehyde ) and little else from 8 hours after application. This was considered to show an auto - oxidation reaction occurred for this product, similar to that observed with the alkyd enamel paint discussed earlier.


Table 16.6 VOC emissions (Brown, 2000 ) from ‘ natural ’ paints (23 ° C, 50% RH , 1.0 h − 1 , 0.5 m 2 m − 3 ).

Paint VOC Emission Factor ( μ g m − 2 h − 1 ) at

8 h 1 d 3 d 14 d

1 Acetone 130 4 < 2 < 2 α , β - Pinene 28 4 < 2 < 2 Camphene < 4 < 2 < 2 < 2 C7 - 10 Alkane 1000 86 14 < 2 Subst. Benzene 150 14 2 < 2 Limonene 2600 180 18 < 2 Nonanal 36 6 2 4 TVOC 4400 420 44 < 20

2 Formaldehyde 98 140 60 (20) a Hexanal 38 64 620 (110) a Nonanal 34 26 52 (24) a TVOC 56 220 840 (400) a

a Measurement at 6 days.

16.3.4 Low - VOC / VOC - Free Paints

International pressure for low - VOC paints and coatings has existed for 10 – 20 years due to the vast quantities of VOCs that paints release into urban (outdoor) atmo-spheres. (Willemse, 1993 ). The typical VOC content of interior water - based latex paints is 180 g l − 1 , although some products with VOC content less than 12 g l − 1 were introduced in the early 1990s (Wicks, Jones and Pappas, 1992a ). Gloss enamel paints are usually alkyd solutions in solvents, and require VOC contents greater than 275 g l − 1 to be functional. However, alkyd enamels with VOC contents of 155 g l − 1 , retaining good fi lm properties, were produced by using reactive diluents which co - react with the alkyds (Wicks, Jones and Pappas, 1992a ). While these reductions may reduce total environmental emissions from paints, the approach is simplistic since health and other environmental issues are ignored. However, in recent years low - VOC and VOC - free paints have become available for lower indoor air impacts, and these will be described in this section. Technical aspects of these changes have been discussed elsewhere (Broek, 1993 ).

Chang et al . (1999) evaluated four USA low - VOC paints in small environmental chambers, in comparison with a conventional water - based paint. All low - VOC paints exhibited TVOC emissions over a 50 - hour test period that were at least an order of magnitude lower than that of the conventional paint. VOCs emitted from the low - VOC paints were ethylene glycol, 1,2 - propylene glycol, dipropylene glycol,


diethylene glycol butyl ether , Texanol ® , and several aldehydes including formal-dehyde and acetaldehyde. Two of the low - VOC paints exhibited large formaldehyde emissions (chamber air concentrations peaking at 3200 and 5500 μ g m − 3 , equiva-lent to EFs of 3200 and 5500 μ g m − 2 h − 1 ) that persisted for more than 50 hours (one paint still exhibiting a concentration of 220 μ g m − 3 ). Since such products were marketed as allowing immediate re - occupation of the building after painting, the formaldehyde emission was concluded to present a potential hazard to occupants.

Lundgren et al . (1999) evaluated the fi eld exposures of painters to a low - VOC and a conventional water - based paints in Sweden. They found that exposures to TVOC, 1,2 - propylene glycol, acetaldehyde and ammonia were 3 – 12 times lower for the low - VOC paint, but exposure to formaldehyde (160 – 180 μ g m − 3 ) was 3 – 4 times higher. Brown (2000) reported that emissions from four Australian ‘ VOC - free ’ paints included typical aromatic VOCs, dibutyl ether, ethylene glycol butyl ether , diethylene glycol butyl ether and Texanol ® , though at order of magnitude reduced levels c.f. conventional water - based paints, as well as formaldehyde and benzaldehyde .

Clearly, a better defi nition for ‘ low - or zero - emission ’ paints is needed.

16.4 Floor Covering Systems

Floor covering systems consist of the structural fl oor and the covering (such as carpet/underlay or resilient fl ooring [PVC, linoleum, rubber, polyolefi n]) applied to the fl oor, in some cases with adhesive bonding. Floor covering systems present a large interior area in all buildings and have an important potential impact on indoor air pollutants. Carpet systems with low emissions have received much attention for over a decade in the USA, and recently the German Federal Environ-mental Agency (Umweltbundesamt) created an environmental label (RAL UZ 113, Blauer Engel, Blue Angel) for low - emission fl ooring adhesives and materials, with emission testing at 3 and 28 days and a health - related evaluation of VOC emissions according to the AgBB (Committee for the Health - Related Evaluation of Building Materials) evaluation scheme (AgBB, 2002 ).

16.4.1 Adhesives

Adhesives are expected to be signifi cant contributors to indoor air pollution where they are used in large surface applications, such as for fl oorcoverings. Yu and Crump (2002) summarize the VOCs identifi ed in emissions from adhesives for carpet, vinyl and parquet tile fl oorcoverings used in the UK. Each adhesive type included a vast array of VOCs, from most classes of VOCs. Glycols, and glycol ethers and esters, and several alcohols were common, showing signifi cant overlap with the VOCs emitted from water - based paints.

16.4 Floor Covering Systems 389

As with paints, adhesives can be water - or solvent - based, or can be ‘ low - emission ’ . Common water - based adhesives are polyvinylacetate ( PVA ), acrylic or rubber latex (natural or synthetic rubber) adhesives. Emission behavior of adhe-sives has received less study than paints. While some of the fi ndings for paint emissions might be relevant for adhesives, in practice adhesive emissions will be more complex due to high solvent levels, high thickness and the interaction with adherends of different physical properties (e.g., permeability). Highly permeable adherends such as carpets may not slow emissions while plastic fl oorcoverings have been found to do so (Wilke, Jann and Br ç dner, 2004 ). This will be important for the common double - bond method of carpet installation, where the underlay is fi rst bonded to the fl oor and then the carpet is bonded to the underlay, both at 100% adhesive coverage. Sheldon et al . (1988) ; Black, Pearson and Work (1991) found that carpet installation adhesives could be far more signifi cant VOC sources than the carpets themselves.

Several early studies of adhesives found they were signifi cant emitters of VOCs, whether water - or solvent - based, such as:

• (in order of decreasing abundance) toluene, styrene, cyclic/branched/normal alkanes (Girman et al ., 1984 );

• toluene, benzene, ethylbenzene, ethyl acetate, styrene (Bayer and Black, 1989 ); • m - xylene, toluene, o - xylene, ethyl benzene (Seifert, Ullrich and Nagel, 1989 ).

Wilke, Jann and Br ç dner (2004) investigated 8 German water - based dispersion adhesives, marketed as very low - emission products for fl oorcovering systems. Generally, the emissions from each adhesive after 28 days consisted of quite dif-ferent VOCs, with only 2 - ethyl - 1 - hexanol being commonly emitted (from 6 of 8 products). Those VOCs listed in Table 16.1 (and highest product EF ( μ g m − 2 h − 1 ) at 28 days) were: 2 - ethyl - 1 - hexanol (45), hexanal (19), 1 - butanol (14), and α - pinene (1.3). Currently in the USA, the Carpet and Rug Institute specifi es formaldehyde and VOC emission limits for carpet adhesives, as presented in Table 16.6 . Note that only 2 - ethyl - 1 - hexanol is specifi ed with a 1 day EF of 3000 μ g m − 2 h − 1 , but it is not possible to compare this value to the above 28 day EF without an understand-ing of the products ’ emission decay behavior.

16.4.2 Carpets and Underlays

Carpets may be constructed as a continuous roll in which the fi ber tufts are bonded into a backing material with a styrene – butadiene rubber ( SBR ) latex adhesive, or as modular tiles held together by a range of binders. VOCs emitted from new SBR - bonded carpets have been found to consist primarily of compounds that originate in the SBR latex (Black, Pearson and Work, 1991, 1993a ; Hawkins, Luedtke and Mitchell, 1992 ; Hodgson, Wooley and Daisey, 1993 ; Bayer and Papa-nicolopoulos, 1990 ). A review of carpet emission studies (Dietert and Hedge, 1996 ) summarized the dominant VOCs emitted from SBR - bonded carpets as follows (in order of decreasing presence in the carpets):


• styrene • 4 - phenylcyclohexene ( 4PC ) • 4 - vinylcyclohexene ( 4VC ) • n - undecane • n - propylbenzene • n - decane • ethylbenzene • ethylene glycol butyl ether • cumene (isopropylbenzene) • 3 - ethyl toluene • toluene • p - xylene.

Many more VOCs may be detectable in lesser quantities. For example, Pleil and Whiton (1990) found that a diverse range of VOCs could be solvent - extracted from carpet, including dichlorobenzene , bis(2 - ethylhexyl)phthalate, triethylphosphate, ε - caprolactam and methylene - bis(4 - isocyanatobenzene). In a more recent study, Wilke, Jann and Br ç dner (2004) evaluated 13 wool and synthetic fi ber carpets in Germany. Those VOCs emitted from 2 or more carpets (and highest EF μ g m − 2 h − 1 ) after 28 days were: 4PC (23), n - tridecane (5), acetic acid (4100), caprolactam (120), n - tetradecane (18), n - heptadecane (6), butylated hydroxyl toluene (BHT) (1), ethylene glycol (25), 1,2 - propylene glycol (50), 2 - ethyl - 1 - hexanol (29), TXIB (7), n - hexadecane (7), benzothiazole (2), and cyclohexanol (2).

The compound 4PC is present in the SBR latex used in carpet manufacture as a contaminant formed in the reaction of styrene and 1,3 - butadiene (Gustafsson, 1992 ). It has a low odor threshold, approximately 3 μ g m − 3 (Dietert and Hedge, 1996 ), and is responsible for much of the odor from new SBR - latex carpets. The compound 4 - vinylcyclohexene (4VC) is a similar contaminant and is also found in new carpet emissions. 4VC is a known carcinogen in animals, inducing ovarian cancer, and is classifi ed as an IARC Group 2B carcinogen ( ‘ possibly carcinogenic to humans ’ ). Limited research has not found the same effect with 4PC (Dietert and Hedge, 1996 ).

Black, Pearson and Work (1991) found no detectable formaldehyde emissions from nineteen SBR - bonded carpet products. In general, the emission of formal-dehyde from carpets appears to be a rare occurrence in published literature.

Some roll carpets use foam polyurethane or PVC backing, and emissions from these may include toluene and acetone (Wolkoff, Nielsen and Hansen, 1990 ) or plasticizer degradation products, such as 2 - ethyl - 1 - hexanol (Gustafsson, 1992 ). Hodgson, Wooley and Daisey (1993) found that such carpets emitted 1,2 - propylene glycol, vinyl acetate, 2,6 - di - t - butyl - 4 - methylphenol and small quantities of formal-dehyde and glycol ethers . Modular carpet tiles have been found to emit trimethyl-pentane (isomer not specifi ed) and acetic acid (Black et al ., 1993b ). Polyurethane foam is also used as a common underlay material. Hugo, Spence and Lickly (2004) investigated the emission of toluylene 2,4 - diisocyanate (TDI) from 3 - day old samples of foam from four US manufacturers in a fl ow through chamber at 37 ° C

16.4 Floor Covering Systems 391

(sample and chamber sizes were not described) but found no TDI was detectable ( < 1.5 μ g m − 3 ) in the chamber exhaust.

There have been concerns about the health impact to building occupants of VOC emissions from new carpets, particularly in the USA (Bayer, 1991 ; Wagner, 1991 ) where a government – industry process of carpet emission assessment was insti-gated. An outcome of this process was a testing and labeling program for new carpets by the Carpet and Rug Institute ( CRI ). New carpets were approved if their emissions at 24 hours after manufacture were less than (Wagner, 1991 ): total VOC (TVOC) 600 μ g m − 2 h − 1 , styrene 400 μ g m − 2 h − 1 , 4PC 100 μ g m − 2 h − 1 and formaldehyde 50 μ g m − 2 h − 1 . Current CRI goals for carpet, underlay and carpet adhesives are presented in Table 16.7 . Note that a goal for 4VC is not included (which is

Table 16.7 US Carpet and Rug Institute emission goals in 2005 for carpet, underlay and adhesives.

Product Type Pollutant Emission Factor at 24 hours μ g m − 2 h − 1

Carpet (green) 4 - phenylcyclohexene 50 Styrene 400 TVOC 500 Formaldehyde 50

Carpet (green plus) a 4 - phenylcyclohexene 5 Styrene 440 Formaldehyde 32 Acetaldehyde 9 Benzene 60 Caprolactam 140 2 - Ethylhexoic acid 50 Naphthalene 9 Nonanal 26 Octanal 14 Toluene 300 Vinyl acetate 200

Underlay 4 - phenylcyclohexene 50 Butylated hydroxytoluene 300 TVOC 1 000 Formaldehyde 50

Adhesive 2 - ethyl - 1 - hexanol 3 000 TVOC 10 000 Formaldehyde 50

a Green plus criteria calculated using area specifi c fl ow rate of 2 m h − 1 (California Department of Health Services, 2004 ); 1 day samples also subject to acceptable 14 day emissions.


surprising since most other emission standards do include 4VC as it is a listed carcinogen).

There have been limited fi eld and controlled exposure studies that evaluated human reactions to carpet emissions. Winfi eld (1987) described a USA primary school where odor, headache, nausea, fatigue and mucosal irritation were reported by a high proportion of the students. Elevated indoor air concentrations of styrene (900 – 4000 μ g m − 3 ) were found, which were believed to be due to a SBR latex - backed carpet which had been installed several years previously. The carpet was removed and the ill - effects were reported to abate. Johnsen et al . (1990) exposed asthma sufferers to a foam rubber - backed carpet in a room chamber. While no clinical effects on lung function were observed, objective eye measurements found that there was a change in tear fi lm quality. This was proposed to result from a degreas-ing effect of lipophilic VOCs, identifi ed as toluene and acetone (Wolkoff, Nielsen and Hansen, 1990 ).

Dietert and Hedge (1996) reviewed the health impacts of VOCs emitted from new SBR - bonded carpets (summarized above), and some of their fi ndings are presented here. They considered that 4VC, which had an occupational exposure standard in the USA (ACGIH, 1997 ) of 440 μ g m − 3 , could exhibit peak exposures from carpets up to 65 μ g m − 3 . Styrene monomer has been reported to have neurotoxic, immunotoxic and mutagenic activity, and has an IARC carcinogen classifi cation 2B (possibly carcinogenic to humans). The occupational exposure standard in the USA (ACGIH, 1997 ) for styrene was 85 000 μ g m − 3 and peak exposures from carpets were estimated to be below 800 μ g m − 3 . Dietert and Hedge (1996) noted that it was not clear that these occupational exposure criteria for 4VC and styrene could be extrapolated to indoor air exposures from carpet emissions. Nielsen, Hansen and Wolkoff (1997) proposed that indoor air goals could be simplistically set to 1/40th of the occupational exposure standards, equivalent to 11 and 2100 μ g m − 3 for 4VC and styrene, respectively. Peak exposure for 4VC from carpets would exceed this goal, though it is considered that quantitative risk assessment should be used to derive an indoor air exposure goal for 4VC.

16.4.3 Plastic Floorcoverings

Plastic fl oorcoverings are generally polymeric materials in which diffusion - based processes are rate - limiting for VOC emissions. Linoleum may be considered an exception here since it has a linseed oil base that autoxidizes, much as alkyd paints described earlier, to release a range of aldehydes. Emission decay from these products is expected to be slow compared to other building products, with long - term emissions being signifi cant. Clausen et al . (1993) considered the role of internal diffusion on VOC emissions from vinyl fl ooring. Using Fick ’ s fi rst law and the observed exponential dependence of the diffusion coeffi cient on VOC concentration in the material, they derived an equation for the emission factor (EF t ) from thick sources which was diffusion - limited:

16.5 Concrete and Plaster Products 393

EF EFt o= ( ) +[ ]− −K t� 1 1

(16.5)

where EF o = emission factor (mg m − 2 h − 1 ) at time t = 0; K = exponential proportion-ality constant (m 3 mg − 1 ) for concentration dependence of diffusivity; and � = thick-ness of the source (m). This model predicts that the emission factor will decrease in proportion to reciprocal time when EF t is large and a stable concentration gradi-ent had been established in the material. They were able to show a fi t of this model to VOC emissions (phenol, cyclohexanone ) from vinyl fl ooring. Jensen et al . (1993) found that the model of Clausen et al . could fi t VOC emissions from linoleum from 24 to 1000 hours, the fi t being better than that for a fi rst order decay model:

EF EFt o= −( )exp k t1 (16.6)

where k 1 is the fi rst order decay rate constant. By contrast, Bremer, Witte and Schneider (1993) found the latter model provided good fi t for VOC emissions from vinyl fl ooring. Christiansson, Yu and Neretnieks (1993) developed a diffusion - limited emission model for VOCs from vinyl fl ooring which predicted late - stage emissions reduced exponentially with time. However, the model was not experi-mentally verifi ed.

It is apparent from such models that VOC emissions from these products, while comparatively low relative to ‘ wet ’ materials, may persist for extended times, espe-cially for larger VOCs with slow diffusion properties. For example, Yu and Crump (1998) reported that new linoleums exhibited TVOC EF ( μ g m − 2 h − 1 ) of 22 – 220 but that a 30 year old sample still exhibited a value of 64 μ g m − 2 h − 1 . By comparison, vinyl fl oorcoverings up to 2 years old exhibited wide variations in TVOC emissions of 91 – 22 000 μ g m − 2 h − 1 , new rubber fl oorings 1400 μ g m − 2 h − 1 and new ‘ soft plastic ’ fl oorings 590 μ g m − 2 h − 1 , new carpet 6 – 410 μ g m − 2 h − 1 . Wilke, Jann and Br ç dner (2004) evaluated 10 fl oorcoverings (5 vinyl, 3 linoleum, 1 rubber, 1 polyolefi n) in Germany. Those VOCs emitted from 2 or more vinyls (and highest EF μ g m − 2 h − 1 ) after 28 days were: n - tetradecane (1), n - pentadecane (1), TXIB (670), diethylene glycol butyl ether (290), N - methyl pyrrolidone (31), 2 - ethyl - 1 - hexanol (13), ethyl hexanoic acid (75), phenol (46). Those VOCs emitted from all 3 linoleums (and highest EF μ g m − 2 h − 1 ) after 28 days were: acetic acid (130), hexanoic acid (28), hexanal (28), propylene glycol ether (7), nonanal (5), decanal (5); the rubber fl oor-ing: benzothiazole (86), benzaldehyde (4), BHT (6), styrene (4), naphthalene (4); the polyolefi n fl ooring: N - methyl pyrrolidone (82).

16.5 Concrete and Plaster Products

While concrete and plaster products are essentially inorganic products, additives used in forming or manufacturing these products may act as VOC and formalde-hyde sources, though little has been published on their emissions.


Concrete may use: plasticizers (e.g., sulphonated melamine and naphthalene formaldehyde condensates), air - entraining agents (alkyl/aryl sulfonate surfac-tants), retarders (hydroxy carboxylic acids such as polyethylene glycol mono - p - nonylphenyl ether) and surface washes (benzalkonium chloride) (RAIA 1997 ). Little has been published on air emissions from concrete additives, their leaching into surface waters appearing to be of greater environmental concern (Ruckstuhl, 2001 ).

Gypsumboard (plasterboard) consists of a core of gypsum plaster laminated with paper on both sides and is probably the most commonly used material for lining of walls and ceilings. In some countries, the plaster core may be treated with sili-cone/solvent mixtures for water resistance, and these products will emit VOCs. Alternatively, a water - based wax emulsion can be used which will eliminate such emissions. The paper used to manufacture the board may have surface treatments that emit volatile organics and ink residues may be present if recycled paper was used. Yu and Crump (2002) reported that the 1 - day TVOC emission factor for plasterboard was 6 – 160 μ g m − 2 h − 1 , and included C 11 - C 17 alkanes , toluene , xylenes , trimethylbenzenes , 1 - butanol , formaldehyde , and hexanal . Colon and Mookherjea (1997) determined that bare gypsumboard emitted formaldehyde and predicted indoor levels from typical residential wall installation of < 25 μ g m − 3 . Lastly, pre-mixed joint compound for gypsumboard may contain fungicides and biocides to prevent spoilage, which are absent from dry mix joint compound (Public Works and Government Services Canada, 2000 ).

16.6 Wood - Based Panels

Wood - based panels are manufactured from wood products and synthetic poly-meric resins, the latter usually based on formaldehyde . They are expected to emit a range of volatile organic pollutants, especially formaldehyde . Formaldehyde emissions from these products have been a concern in many countries for over 20 years (Sundin, 1985 ; CEC, 1990 ; Salthammer, Fuhrmann and Kaufhold, 1995 ). As a result, there are standards in place for measuring this property in most countries. Many European countries classify particleboard products according to their formaldehyde emissions measured in a dynamic environmental room chamber at 23 ° C, 45% RH, a ventilation rate of 1.0 air changes per hour (ACH) and a loading ratio (area of board per volume of room) of 1.0 m − 1 . Emission clas-sifi cations usually require that the formaldehyde concentration at ‘ equilibrium ’ in chambers be equal to or less than 120 μ g m − 3 at room temperature (CEC, 1990 ), an equivalent concentration to many indoor air goals. A recent review of the impact of medium - density fi berboard ( MDF ) use on formaldehyde levels in UK homes concluded that transient irritation could be experienced by some individu-als, but that there was no likelihood of other health effects (Harrison and Brooke, 1999 ). However, it was recommended that exposure to formaldehyde from this source should be reduced where possible. Currently, several more stringent

16.6 Wood-Based Panels 395

product emission classifi cations exist with emissions as low as detection levels for formaldehyde being specifi ed in some circumstances (e.g., Japanese building standards).

Investigation into VOCs emitted from these products is less substantial than investigation into formaldehyde emissions. Nelms, Mason and Tichenor (1986) investigated formaldehyde and VOC emissions from a USA sample of new par-ticleboard in small chamber experiments. The major VOCs emitted were (highest to lowest concentration):

• acetone • benzaldehyde • 2 - propanol • benzene • ethyl methyl ketone ( EMK ).

They observed that the emission factors for formaldehyde , benzaldehyde and pos-sibly EMK were increased as the ratio of ventilation rate to loading ratio (product area per chamber volume) was increased. That is, the pollutant emission rates increased as the ventilation rate was increased or the loading ratio was decreased, indicating that these pollutants were emitted at rates dependent on the pollutant concentrations in the surrounding air.

Leovic (1996) reported that the combined emissions of VOCs as ketones, ethers and aldehydes were generally much in excess of formaldehyde emissions for veneered particleboards. Koontz and Hoag (1995) had a similar fi nding for unfi n-ished and veneered particleboard and MDF. They identifi ed the major VOCs emitted as (highest to lowest):

• acetone • hexanal • pentanal • benzaldehyde • pentanol (isomer not specifi ed) • heptanal • pinene • nonanal • octanol.

Brown (1999b) reported formaldehyde and VOC emissions from new, unfi nished particleboard and MDF (both using urea formaldehyde resins) in Australia. Form-aldehyde emissions over the fi rst three weeks exhibited fi rst - order decay behavior that predicted little to no formaldehyde emission after 6 months. However, further emission measurements at ∼ 8 months showed the products still emitted formal-dehyde at approximately one - half the new product rate (also further unpublished measurement at 2 years showed the same emission rate as at 8 months). It was concluded that the wood - based panels emitted formaldehyde by a double - exponen-tial model, the early - to late - term emissions including the free formaldehyde in the products but the long - term emissions consisting of only the formaldehyde


released by hydrolytic degradation of resins in the boards, which would occur over the lifetime of the product.

It was found that the new MDF specimen emitted no VOC at emission factors greater than 8 μ g m − 2 h − 1 or TVOC concentrations greater than 25 – 50 μ g m − 2 h − 1 in the fi rst two weeks of assessment. However, when retested after nine to 10 months it emitted hexanal at 80 – 100 μ g m − 2 h − 1 . No explanation is available for this change in emission behavior, though subsequent (unpublished) assessment of MDF from another Australian manufacturer found measurable emissions, similar to those described below for particleboard. The particleboard emitted signifi cantly greater amounts of VOCs than MDF, similar to those found by Koontz and Hoag (1995) , although an extra compound (methanol) was the major VOC emitted (probably formed by the Cannizzaro reaction; Meyer, 1979 ) at 1300 μ g m − 2 h − 1 . All of these emissions, except hexanal and nonanal , decayed rapidly by fi rst order decay over the fi rst two weeks to attain low levels similar to those found for MDF. Hexanal emission remained approximately constant over the two weeks, but had decreased by 6 to 7 months to much lower levels than were observed for MDF after such a period. Yang et al . (2001) investigated models to numerically simulate the VOCs emitted from two new Canadian particleboards. The major compounds identifi ed were the same for the boards: hexanal , α - pinene , camphene, and limonene , but the emissions were higher for the board containing a higher content of formalde-hyde scavenger.

16.7 Natural Wood

Natural woods are important from the indoor air perspective of their use in cultural heritage organizations to house their collections in semi - sealed display cases or storage cabinets. It is well recognized that the materials of construction of display and storage cases, especially some wood products, can emit a range of volatile organic air pollutants that physically degrade cultural items. Of most concern is the damage caused to lead bronzes, calcareous materials (limestones, shells) and glass, sometimes in periods as little as a few months. The pollutants considered to cause greatest damage (Bradley and Thickett, 1998 ) are (in decreasing order): acetic acid, formic acid and formaldehyde . The organic acids are emitted from some raw timbers, such as oak, Douglas fi r and birch and from wood - based panels using these timbers, and may continue to be emitted for many decades.

A summary of emission assessment for European woods and Australian woods and wood - based panels (as above) is presented in Table 16.8 . Note that:

• There appeared to be no advantage in using zero - formaldehyde MDF since it exhibited greater emissions of the organic acids than conventional MDF.

• Lower formaldehyde emissions were found for the plywoods (all exterior grades) than for particleboard and MDF, but organic acid emissions still occurred, especially for the hardwood - based plywood (black butt) since such woods are

16.8 Ovens and Heaters 397

Table 16.8 Emission of organic pollutants from European woods (Nordic Wood, 1998 ) and Australian woods and wood - based panels (Brown, 1999b ); Brown unpublished) ( – no data available).

Product (Europe/Australia)

Maximum emission factor ( μ g m − 2 h − 1 )

Acetic acid

Formic acid

Formal - dehyde

Acetal - dehyde

Hexanal Hexanoic acid

Pine 1 – – – 200 60 – Pine 2 20 – – – 40 40 Spruce 1 – – – 370 30 – Spruce 2 15 – – – 4 3 Birch – – – 10 50 – Beech 7 – – – 17 8 Oak on Spruce – – 8 36 9 – Birch plywood 300 – 5 20 40 5 Laminated Birch plywood – – 9 13 – – Particleboard – – 440 – 50 – MDF1 – – 370 – 80 – MDF2 60 < 20 260 – 14 – Zero - formaldehyde MDF 200 80 < 10 – – – Hoop Pine Plywood 1 < 20 < 20 10 – – – Hoop Pine Plywood 2 700 600 < 10 – – – Black butt plywood 850 < 50 110 – – – Fiji hardwood plywood < 50 < 50 20 – – – Solid Sugi cedar < 50 120 < 10 – – – Solid hoop pine 400 140 < 10 – – – Acid - free cardboard < 50 < 50 < 10 – – –

expected to contain higher contents of acetylated hemicellulose, the probable source of acetic acid emissions (Rydholm, 1965 ).

• Hoop pine plywood, a material used by some museum conservators in Australia, exhibited a large variation in emissions: one sample was virtually nonemitting, the other was high emitting for organic acids, with a solid hoop pine sample confi rming that this wood can emit organic acids.

• It is diffi cult to recommend any wood product for museum cases without a more extensive assessment of these materials.

16.8 Ovens and Heaters

Brown, Cheng and Mahoney (2005) found that new electric ovens were signifi cant sources of formaldehyde and VOCs from the ‘ burn - off ’ of phenol - formaldehyde -


Table 16.9 Volatile organics emitted from fi rst operation at 250 ° C of new electric oven (chamber volume 33.5 m 3 , 23 ° C, 50% RH , 1.1 air change per hour).

Pollutant Concentration ( μ g m − 3 )

1 h 4 h 1 d 2 d

Formaldehyde 2400 850 110 37 Acetone 520 140 18 14 Butanal/MEK 160 23 4 3 Benzene 30 13 1 1 1 - Butanol 72 28 < 1 < 1 Acetic acid 120 68 7 2 3 - Methylbutanal 92 28 3 2 Toluene 12 6 < 1 < 1 Hexanal 100 29 2 2 Heptanal 120 34 4 2 Phenol 110 59 27 12 Octanal 96 34 4 4 Nonanal 82 30 4 2 Decamethylcyclopentasiloxane 560 140 8.9 5.4 3 - Hydroxy - 2 - methylbenzaldehyde 47 11 < 1 < 1 2,4 - Dichlorophenol 93 32 3 2 Decanal 56 15 1 1 Decamethylpentasiloxane 3700 880 28 15 Dodecamethylpentasiloxane 2100 900 34 18 Tetradecamethylhexasiloxane 750 570 41 16 Hexadecamethylheptasiloxane 230 220 55 24 TVOC 5300 1800 160 97

bonded mineral fi ber insulation, residual oils and silicone wiring used in oven construction. The organics released from a typical new oven operating continually for 48 hours are summarized in Table 16.9 .

In other unpublished work by CSIRO ’ s laboratory, emissions were evaluated from an electric heater panel that was designed to replace offi ce ceiling tiles. Spe-cifi c construction details were unknown other than that the panel exterior was powder - coated steel. Emissions were dominated by 1 - butanol and it was assumed that this originated from the powder coating, as may be the case with the oven above.

Heaters that are not vented to the exterior have the potential to act as indoor pollutant sources, especially gas heaters. While emissions of carbon monoxide and nitrogen dioxide have received investigation for more than a decade, it has been found that ‘ low - NOx ’ unfl ued gas heaters can act as sources of formaldehyde to indoor air (Brown, Mahoney and Cheng, 2004 ). Table 16.10 summarizes volatile organic emissions from an unfl ued gas heater (A) where it was apparent


that modifi cation to the combustion process to reduce NOx emissions resulted in less effi cient combustion and the byproducts formaldehyde and several VVOCs (possibly contaminants in natural gas escaping during heater on/off operation). Much lower organic emissions occurred from a heater by another manufacturer (B).


It is clear that VOC and formaldehyde emissions from building products are diverse and complex processes and that there is a growing understanding of these processes. This chapter has provided a snapshot of this situation, and has tried to identify important links between VOC emissions over time, especially those of concern to occupant health and well - being, and the key sources of these emissions. Initially, emissions were determined by empirical measurement over limited time periods and under limited physical conditions, the latter simulating conditions of use in practice. We have seen the development of semi - empirical emission models for building products over the last decade, linked to product use/application, build-ing physical conditions and surface sink effects. These have allowed modeling of indoor air concentrations and (ideally) have underpinned the pollutant emission limits for building products. Notably, the latter are generally restricted to mid - to late - term emissions, not long - term.

Table 16.10 Volatile organic emissions from two unfl ued gas heaters operating at 5 – 7 MJ h − 1 in 33.5 m 3 room chamber at 25 ° C, 50% RH and 2 air changes per hour.

Pollutant Chamber Concentration ( μ g m − 3 )

Heater A Heater B

Formaldehyde 130 < 10 Low boiling alkanes 27 6 Isobutane 1 3 Butane/Methanol 11 < 1 2 - Methylbutane 1 < 1 Acetone/3 - methylbutanal 12 3 Methylcyclopentane < 1 < 1 Acetic acid < 2 3 Benzene < 1 < 1 Benzaldehyde < 1 2 Phenol < 1 1 Other VOCs < 1 < 1


The most recent research is directed at a more fundamental understanding of emissions based on the physical - chemical processes that occur in building materi-als. Such understanding has potential to lead to redundancy of the semi - empirical measurements now used in guidelines and regulations, emission limits instead relying only on the compositional analysis of the initial product. However, it must be remembered that building products are manufactured products for which technological advances will change volatile and nonvolatile components, product use and other factors, all which may present moving targets to fundamental modeling.

The aim of all this research is to reduce the emissions to levels at which the health and well - being of building occupants are not impaired – this is the key outcome sought by all indoor air research. Future research needs to be more focused on attaining such an outcome by ensuring that research fi ndings can be implemented by manufacturers and by building, health and environmental regula-tors. Several countries have now implemented regulations for building product emissions; the success of these will depend on how well they have been based on rigorous emission research that targets key pollutants, key building products, and appropriate occupancy factors (especially time) such that indoor air exposures are reduced to acceptable levels.

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Emission of VOC s and SVOC s from Electronic Devices and Offi ce Equipment Tobias Schripp , and Michael Wensing

17

17.1 Introduction

Nowadays the electronic appliances used for entertainment, telecommunications and data processing are widespread in daily life. Typical examples include televi-sions, video recorders, hi - fi systems and fax machines, not to mention computers with their peripherals such as monitors and printers, scanners and copiers. These devices are predominantly made of polymeric components and materials which might contain additives, such as fl ame retardants and plasticizers (Wensing, Uhde and Salthammer, 2005 ) to obtain specifi c desired properties. In addition, there will also be chemical residues from production and processing aids, such as residual monomers and solvents. Especially under operating conditions these compounds can be released from electronic equipment into indoor air due to the heating - up of the device interior. In many cases, such emissions can be monitored via simple odor tests (Walpot, 1996 ).

In contrast to other sources of emissions indoors, such as building materials and furnishings, electronic devices are ‘ active ’ emission sources. Firstly, they consume electrical power which can result in heat being generated inside the device, which can in turn result in the generation of higher levels of emissions. As an example, Figure 17.1 shows a thermographic image of the heat distribution in the main printed circuit board of a television set of the cathode ray tube type (T Ü V NORD, 1999 ). It can clearly be seen that there are hot areas inside a device which favor the appearance of VOC and SVOC emissions from the various com-ponents when the device is in operation. Secondly, in case of equipment such as printers and copiers (hard - copy devices), in addition to emissions from the device itself the constantly recurrent emissions from paper, toner and ink ( ‘ consum-ables ’ ) also contribute to the total emissions. A large number of studies have been devoted to emissions from electronic devices. Black and Worthan (1999) have described the VOC/TVOC, particle and, ozone emissions of laser printers, dry - process photocopiers and personal computers. In the same year Brown (1999) published emissions data for VOCs, formaldehyde, respirable particles, ozone and


405

406 17 Emission of VOCs and SVOCs from Electronic Devices and Offi ce Equipment

nitrogen dioxide from dry - process photocopiers, while one study by Wolkoff (1999) dealt with photocopiers and indoor air pollution and another by Wensing (1999) reported on the VOC and SVOC emissions of televisions and video recorders. Table 17.1 reproduces a list taken from Wensing ’ s study of the various VOCs/SVOCs which were identifi ed qualitatively by GC – MS and in part also quantifi ed during operational testing of 10 televisions and 10 video recorders fresh off the production line (T Ü V NORD, 1999 ; Wensing, 1999 ). This list essentially contains substances which can be emitted by the materials making up the casing and from the electronic components. Since these materials and components are used in a great variety of ways not only in television sets and video recorders but also in the most varied types of electronic equipment, this emissions spectrum should also be typical for other devices. Additional quantitative data for individual VOCs as well as for formaldehyde and acetaldehyde emissions from device casing materials and electronic components may be found in Funaki et al . (2003) , Nakagawa et al . (2003) , M ö ller et al . (2004) , and Wensing (2004) .

Data presenting emissions from hard - copy devices in operation may be found in, for example, Wensing et al . (2002a) , Rockstroh et al . (2003) , Rockstroh et al . (2005) , Jann et al . (2005) , Jann and Wilke (2006) and Kagi et al . (2007) . Some of the emissions are suspected of being the source of irritations and adverse health effects (Wargocki et al ., 2003 ; Bak ó - Bir ó et al ., 2004 ; Wolkoff et al ., 2006 ). The emissions of printers and copiers in particular are the focus of discussion in Smola, Georg and Hohensee (2002) ; Ewers and Nowak (2006) ; Roller (2006) ; Gminski and Mersch - Sundermann (2006) . Alongside VOC and SVOC emissions, another topic of high interest is the release of the ultra - fi ne particles (particle size

Figure 17.1 Thermographic image of the heat distribution of the cathode ray tube board of a television set (T Ü V NORD, 1999 ).

17.1 Introduction 407

Table 17.1 VOC s / SVOC s (selection) identifi ed by GC – MS during emission tests of TV sets and VCR (Wensing, 1999 ).

Compound classes Selected compounds

Aliphatic hydrocarbons n - alkanes, isoalkanes, cycloalkanes (C 10 – C 35 )

Aromatic hydrocarbons Benzene, toluene, ethylbenzene, propylbenzene, 1,3,5 - trimethylbenzene, 1,2,4 - trimethylbenzene, 1,2,3 - trimethylbenzene, tert - butylbenzene, o,m,p - xylene, p - cymene, styrene, alpha - methylstyrene, naphthalene, methylnaphthalene

Phenols/cresols Phenol, 2,5 - dimethylphenol, 3,5 - dimethylphenol, 2,6 - di - tert - butyl - 4 - methylphenol, o,m,p - cresol, p - tetramethylbutylphenol, nonylphenols, 2,6 - di - tert - butyl - 4 - methoxymethylphenol, 2,6 - di - tert - butyl - 4 - methylphenol, 4,4 ′ - butyliden - bis - (6 - tert - butyl) - m - cresol, C - 8 - alkylphenols, C - 9 - alkylphenols

Halogenated hydrocarbons Dichloromethane, trichloroethene, tetrachloroethene, tetrachloromethane, trichloromethane, 2 - chlorophenol, 3 - bromoheptane, pentabromotoluene, hexabromobenzene, polychlorinated biphenyls (PCB)

aldehydes butanal, pentanal, hexanal, heptanal, octanal, nonanal, decanal, undecanal, dodecanal, furfural, 2 - methyl - 2 - pentenal, 2 - ethyl - 1 - hexanal, cis - 2 - heptenal, 2 - decanal, 4 - hydroxybenzaldehyde, 4 - methoxybenzaldehyde, dehydroabietic aldehyde

Ketones 2 - butanone, acetophenone, benzophenone, 2,2 - dimethoxy - 2 - phenyl acetophenone

Acids Acetic acid, 2 - ethylhexanoic acid, methacrylic acid, benzoic acid, dehydroabietic acid, p - tert - butylbenzoic acid

Fatty acid esters Ethyl acetate, methyl methacrylate, trimethylolpropane trimethacrylate, decyl methacrylate, 2 - hydroxyethyl methacrylate, ethoxy ethyl acetate, methyl palmitate, ethyl palmitate, isopropyl palmitate, butyl palmitate, isobutyl palmitate, methyl margarate, butyl margarate, methyl stearate, ethyl stearate, butyl stearate, isobutyl stearate, methyl oleate, bis(2 - ethylhexyl) adipate, bis(2 - ethylhexyl) sebacate, methyl linoleate, methyl laurate, octyl laurate, methyl myristate, isobutyl myristate, butyl myristate, dihydroisopimaric acid methyl ester

Alcohols 1 - butanol, 1 - pentanol, 1 - hexanol, 2 - ethyl - 1 - hexanol, 1 - heptanol, 1 - octanol, 1 - decanol, 1 - methoxy - 2 - propanol, butoxyethanol, ethoxyethanol, 2 - (2 - butoxyethoxy) ethanol, benzyl alcohol, 1 - dodecanol, 1 - tetradecanol, 1 - pentadecanol, 1 - hexadecanol, 1 - octadecanol

Si - compounds Trimethylsilanol, hexamethylcyclotrisiloxane, octamethylcyclotetrasiloxane, decamethylcyclopentasiloxane, dodecamethylpentasiloxane


Compound classes Selected compounds

Phthalates Dimethyl phthalate, diethyl phthalate, diisobutyl phthalate, dibutyl phthalate, dicyclohexyl phthalate, bis(2 - ethylhexyl) phthalate, (2 - ethylhexyl) methylphthalate

phosphates phosphoric acid trimethyl ester, phosphoric acid triethyl ester, triphenyl phosphate, tris(chloropropyl) phosphate, tris(2 - chloroethyl) phosphate, diphenyl cresyl phosphate

Terpenes Alpha - pinene, delta - 3 - carene, limonene, longifolene

Others 1,4 - dioxane, 1,3,5 - trioxane, terahydrofuran, 2 - butyl tetrahydrofuran, N,N - dimethylformamide, N,N - dimethylacetamide, epsilon - caprolactam, benzothiazole, biphenyl, pyrazine, 2 - butanone oxime, 1 - methyl - 2 - pyrrolidone, N,N - dimethyl - aniline, N,N - dimethyl - p - toluidine, 1 - methyl - 7 - isopropyl phenanthrene, dibenzyl - amine, hexamethylenetetramine, squalene, 2 - ethylanthraquinone


< 0.1 μ m) (Bake and Moriske, 2006 ; Seeger et al ., 2006 ; Uhde, He and Wensing, 2006 ; Wensing et al ., 2006, 2008 ; Kagi et al ., 2007 ). Regarding SVOC emissions, Carlsson, Nilsson and Ö stman (2000) described emissions of triphenyl phosphate from computer video display units. Sj ö din et al . (2001) examined the air in work-places – a plant engaged in recycling electronic goods, a factory assembling printed circuit boards, a computer repair facility, and offi ces equipped with computers – for phosphor - organic compound s ( POC ) and brominated fl ame retardants. With regard to these classes of chemical compounds, emission test chamber measure-ments described by Kemmlein, Hahn and Jann (2003) , M ö ller et al . (2004) , Wensing (2004) , and Wensing, Uhde and Salthammer (2005) provide further emissions data for POC and phthalates emitted from casing materials and electronic components. In the case of one television set, emissions of short - chain chlorinated paraffi n s ( CP s) were detected (Wensing, 2003 ).

17.2 Test Procedures

With the aid of emission test chamber measurements (see Chapter 5 ) emissions from electronic devices can be determined without any environmental infl uences affecting results. Such investigations aim at determining device - specifi c emission rates under standardized environmental conditions. Using these emission rates it is possible to make comparisons between the emission characteristics of different devices not only qualitatively (the composition of the emissions) but also quanti-

17.2 Test Procedures 409

tatively (absolute quantity of emissions). The results of investigations of this kind provide engineers working on the development of new devices with valuable information, allowing them to identify ‘ toxicologically critical ’ emissions at an early stage and avoid them in future by making the appropriate selection of materi-als or by making modifi cations during the course of production. Basing on the emission analysis, low - emitting devices can be awarded an environmental label such as the Blue Angel (RAL, 2006 ). The results of emissions testing can also be used for calculating the theoretical indoor air concentrations to be expected in for a model room (e.g., Wensing et al ., 2002a ; Wensing, 2004 ).

In principle, the same emission test chambers developed for testing building products (DIN EN ISO 16000 - 9, 2006 ), can also be used in testing electronic devices. Figure 17.2 presents a view into an open 1 m 3 test chamber with walls made of electro - polished stainless steel in which a television is running for emis-sions testing. DIN EN ISO 16000 - 9 (2006) describes fundamental quality require-ments which generally need to be observed in the case of emission test chamber measurements. Due to the fact that electronic devices are ‘ active ’ emissions sources special aspects need to be taken into consideration when their emissions are studied and during the course of testing: these include duration of testing, sampling times and testing parameters as described in ECMA 328 (ECMA, 2007 ). Regarding the emission testing of electronic devices a distinction is drawn between equipment under test ( EUT ) with and without the use of consumables. The third edition of ECMA (2007) is also fully aligned with the fi rst edition of ISO/IEC (2007) and describes the determination of emission rates of VOC (including TVOC), ozone and particulate matter. On the basis of these test parameters the Blue Angel environment award (RAL, 2006 ) is given in Germany to hard - copy devices (printer, copier and multi - function devices) with particularly low emission rates. The corresponding test protocol for emissions testing is also an integral part of ECMA (2007) .

Figure 17.2 View into 1 m 3 emission test chamber with test subject (TV set) and fogging plate (T Ü V NORD, 1999 ).


VOC and SVOC emissions generally occur in the following three different operating states:

• Pre - operating phase: The phase in which the EUT is connected to an electrical supply and which may include warming - up and energy - saving modes; this is the phase before the EUT is able to enter the operating phase.

• Operating phase: The phase in which the EUT is performing its intended functions.

• Post - operating phase: The phase following the operating phase. The post - operating phase may include energy - saving modes.

During emissions testing following the ECMA 328 scheme a device runs through all three phases in turn. In the case of equipment running with use of consum-ables, the constantly recurrent emissions from the consumables are added in the operating phase to the casing and materials emissions. Particularly in the case of VOC emissions testing of hard - copy devices there may be a complex superposition of the various emissions during the pre - operating, operating and post - operating phases (in this connection see Figure 17.3 , which is based on a model with the following assumptions [ECMA, 2007 ]):

• the emission rates are constant; • the emission rates during the post - operating and operating phases are equal; • no energy - saving modes apply during the pre - and post - operating phases.

After switching on (pre - operating phase), the emissions of the device increase and under the boundary conditions of the test (air exchange rate) an equilibrium con-centration C pre establishes on top of the background concentration C bg of the emission test chamber. During the operating phase there occur the emissions from consumables and heated electronic components. Once the operation has

Figure 17.3 Concentration curve for a printing test in an emission test chamber. Emissions without printing are shown in dark gray, printing emissions in light gray (diagram taken from ECMA, 2007 ).


Figure 17.4 Concentration curve for a printing test in an emission test chamber, model as per RAL UZ 122. Emissions without printing are shown in dark gray, printing emissions in light gray (diagram taken from ECMA, 2007 ).

been completed, concentrations in the emission test chamber drop according to the air exchange and the device is in the post - operating phase. Regarding the emission behavior this phase is in principle identical to the pre - operating phase. The contribution to the total emission by the pre - and post - operating phases is shown in dark gray in Figure 17.3 . The light gray area corresponds to the emis-sions originating in printing mode (operating phase).

Similar to Figure 17.3 , Figure 17.4 shows the concentration development during a printing operation. In this case, the emission model is identical to the one used in RAL UZ 122. The main difference is the addition of the emissions from the pre - operating and post - operating phases to the emissions of the printing phase. According to the RAL UZ 122 (2006) test protocol, air samples are taken at the end of the pre - operating phase and at the beginning of the printing phase in order to determine VOC (see Figure 17.4 ). The emission rate for the pre - operating phase can be calculated from the concentration of the sample taken over the last twenty minutes of the one - hour phase using the equations:

SERB B B= ⋅ ⋅C n V (17.1)

with

Cm

VB

VOC

P

B= (17.2)


where

C B VOC concentration ( μ g m − 3 ) during the pre - operating phase SER B VOC emission rate ( μ g h − 1 ) during the pre - operation phase m VOCB

Analyzed mass ( μ g) of VOC during the pre - operating phase n B Air exchange rate (h − 1 ) during the pre - operating phase V Volume of test chamber (m 3 ) V P Sample volume (m 3 ) during the pre - operating phase

The emission rate during the printing phase is given by Equation 17.3 . This cal-culation considers the amount of VOC emitted during the printing and post - operating phase. The length of the post - operating phase is set to one air exchange of the chamber volume.

SERSER

DN

VOC

PDN G B DN G

DN D

DN

DN G D=

⋅ ⋅ ⋅ − ⋅ ⋅

⋅ − +− ⋅ −

m

Vn V t n t

n t e en t t

2

( ) −− ⋅n tDN G (17.3)

where

SER DN VOC emission rate ( μ g h − 1 ) determined from the printing and post - operating phases

SER B VOC emission rate ( μ g h − 1 ) determined from the pre - operating phase m VOCDN

Analyzed mass ( μ g) of VOC during the printing phase and post - operat-ing phases

n DN Air exchange rate (h − 1 ) during the printing and post - operating phases t D Absolute printing or copying time (h) t G Overall sampling time (h) V Volume of test chamber (m 3 ) V P Sample volume (m 3 ) during the printing and post - operating phases

As has been mentioned earlier, especially low - emitting hard - copy devices can be awarded the Blue Angel in Germany if the award criteria (RAL, 2006 ) are satisfi ed. To check this, test pages are printed out in monochrome black/white mode with coverage of 5% and in color mode with coverage of 20%. Table 17.2 presents the corresponding limit requirements to be observed and a distinction is also drawn between so - called tabletop and fl oor - mounted devices. Ozone emissions are measured with continuously recording measuring instruments while dust emissions are determined gravimetrically by fi lter measurement. For detailed information, see ECMA (2007) , ISO/IEC 28369 (2007) and RAL UZ 122 (2006) . To monitor the VOC emissions, sampling from the test chamber air is carried out by active sampling using Tenax TA as sorbent. This collection phase is then thermally desorbed and evaluated by GC – MS (DIN ISO 16000 - 6, 2004 ).

With the exception of POC (VDI 4301 - 5, 2009 ), no special emission chamber measurements have been standardized for investigating the SVOC emissions (e.g.,


plasticizers, fl ame retardants) of electronic devices until now. In the corresponding tests the same analytic methods can be used as were developed for room air mea-surements (Wensing, Schulze and Salthammer, 2002b ). When the conditions (e.g., duration of the test) of an SVOC emission test chamber measurement are selected, consideration should always be given to the fi ndings from investigations into other materials (e.g., Uhde et al ., 2001 ; Horn, Jann and Wilke, 2003 ). Due to sorption of SVOC to the test chamber walls (sink effect) the time to reach equilib-rium concentration in the chamber can be considerably higher compared with VOC measurements. In addition to the SVOC sampling on a polymer collection phase (e.g., PU foam or XAD - 2) it is also possible to use a cooled ‘ fogging plate ’ (visible in the foreground in Figure 17.2 ) for the deposition of SVOC emissions from electronic devices (T Ü V NORD, 1999 ; Wensing, 1999 ; Wensing, Uhde and Salthammer, 2005 ). With this technique, which was fi rst used in the testing of automobile interiors (Bauhof et al ., 1996 ) and subsequently with other products as well (Uhde et al ., 2001 ), the SVOC are deposited on a cooled surface, eluted with an organic solvent, and then analyzed by GC – MS.

The evaluation of emission test chamber results aiming at the estimation of exposure against emitted compounds is usually carried out by calculating the concentrations in a model room. The parameters of this room are defi ned in DIN EN ISO 16000 - 9 (2006) including a volume of 17.4 m 3 and an air exchange of 0.5 h − 1 . By determining the emission rates of an electronic device it is possible to estimate the pollutant concentrations under real conditions and evaluating them under aspects of the recommendations for a good IAQ (Wensing et al ., 2002a ; Wensing, 2004 ).

Table 17.2 Limit requirements of the Blue Angel for particularly low - emission hard - copy devices ( RAL , 2006 ).

Monochrome (b/w) values in (mg h − 1 )

Color values in (mg h − 1 )

Pre - operation phase TVOC 1 a 2 b

1 a 2 b

Sum of pre - operation and operation phase TVOC 10 18 benzene 0.05 0.05 styrene 1.0 1.8 ozone 1.5 3.0 dust 4.0 4.0

a Tabletop - device. b Floor - mounted device.


17.3 VOC and SVOC Emissions from Various Devices

17.3.1 Printers and Copiers

The printing process of a copier and laser printers bases on the electro - photo-graphic principle developed in 1938 (Schein, 1992 ). After scanning the original image the print data is transferred to a character generator that forms a latent image on a photoconductor drum. Then toner is added to this unit to get a toned image that can be transferred on the paper. During the fusing step the toner is fi xed on the paper by use of high temperatures (150 – 220 ° C). For the application of the toner some additives like silica and wax are necessary. Beside these separat-ing agents ‘ carriers ’ consisting of iron globules are used. These particles feature diameters between 35 and 300 μ m and are utilized as charge control substances (total content in the toner: 1 – 3%). Other toner ingredients are resin (80 – 90%) and integrated color pigments (5 – 15%) like carbon black. Laser printers and copiers are a known source for ozone, dust and (S)VOC (e.g., Wolkoff et al ., 1993 ; Black and Worthan, 1999 ; Wensing et al ., 2002a ; Rockstroh et al ., 2003 etc ). There are several health - related impacts of these pollutants. Ozone is a reactive and toxic gas that is able to form secondary emissions volatiles and particulates by reaction with VOC in air (Weschler, 2000 ). The emission of UFP from laser printers will be discussed in the last chapter of this article. Ozone dominated the specifi c odor of a working laser printer in the past. In contrast to older laser printers modern devices emit only low amounts of ozone due to constructive improvements or the use of fi lter assemblies. In some cases the emission of ozone is not detectable at all.

Emissions during application are typically connected to the use of the above mentioned consumables like toner, ink and paper. Especially for hard - copy devices high differences between the background emissions from the casing in idle mode and the emissions during application exist. This is extensively observed for table-top and fl oor - mounted devices (Jann et al ., 2005 ; Jann and Wilke, 2006 ). Some results for printers in idling and working mode are displayed in Table 17.3 . By measuring these two operation types it is possible to distinguish between perma-nent and short period emissions. From the results it can be derived that the use of toner is the most important aspect for VOC emission estimation. In most cases toner consists of small particles (2 – 10 μ m) of a thermoplastic polymer (e.g., styrol - acrylat - copolymer) that is endowed with pigments (Ewers and Nowak, 2006 ). Toner is alleged to be widely biological inert because it is not water soluble. Dusts with such a characteristic are called ‘ low toxicity poorly soluble particle s ’ ( LTPSP ). The polymer that is necessary to fi x the pigments on the paper can be a source for organic emissions but also contaminations with heavy metals might occur. Tin - organic compounds are used as catalysts and stabilizers during the production of resin and are sometimes detected beside other heavy metals in the fi nal product. This broad range of organic and inorganic compounds which can be released from

17.3 V

OC

and SVO

C Em

issions from V

arious Devices

415

Table 17.3 Emission rates of different electronic devices.

Emission rate ( μ g unit − 1 h − 1 )

TV set a VCR a) Remote Control b

Cellular phone b

Tabletop printers (P90) c

CRT - PC d TFT - PC d PC e

Idle Print

Number of devices 10 10 1 1 13 4 4 6 f TVOC 189 – 2036 116 – 391 5.5 7.9 1422 11 647 180 ± 56 113 ± 31.8 468.6 1 - Butanol – – – < 1 7 321 – – – 2 - Butoxyethanol 1.6 – 31.4 < 0.1 – 2.3 – – – – – – – 2 - Ethylhexanol – – – – 15 260 – – 19.6 Acetaldehyde – – – – – – 3.6 ± 1.2 1.5 ± 1.2 – Aromatic hydrocarbons – – – – – – 103 ± 36.1 32 ± 14 – Benzaldehyde – – – – 3.8 189 – – – Benzene 0.3 – 7.1 < 0.1 – 1.2 – 5 28.6 x x – BHT – – – – 50 1460 – – – Decane – – < 1 – – 54 – – 11.6 Dichloromethane < 0.2 – 2.1 < 0.2 – 1.0 – – – – – – – Dodecane – – 1.5 < 1 5 566 – – 22.3 Ethylbenzene – – – – 79 1375 x x – Formaldehyde < 0.5 – 9.9 8.0 – 40 – – – – 12.8 ± 3.6 9.7 ± 3.2 5.2 Hexadecane – – < 1 – – – – – – Hexaldehyde – – – – – – 4.6 ± 0.5 2.7 ± 0.3 – m/p/o - Cresols 0.9 – 19 1.0 – 7.6 – – – – – – – n - Butyraldehyde – – – – – – 1.4 ± 0.9 – – n - Hexane – – < 1 – 100 54 – – – n - Nitrosodibutylamine < 0.1 – 0.2 < 0.1 – – – – – – –

416

17 Emission of V

OC

s and SVO

Cs from

Electronic Devices and O

ffi ce Equipment

Emission rate ( μ g unit − 1 h − 1 )

TV set a VCR a) Remote Control b

Cellular phone b

Tabletop printers (P90) c

CRT - PC d TFT - PC d PC e

Idle Print

Phenol 12 – 326 18 – 65 – – 6 259 – – 63 Propionaldehyde – – – – – – 0.5 ± 0.3 – – Styrene 1 – 7.6 0.2 – 6.6 – – 16 446 x x 7.6 Tetrachloroethene < 0.2 – 2.6 < 0.2 – 2.1 – – – – – – – Tetradecane – – < 1 < 1 5 296 – – – Toluene 1 – 203 2.5 – 67 < 1 < 1 550 1414 x x 47 Trichloroethene < 0.2 – 59 < 0.2 – 4.1 – – – – – – – Trimethylbenzene – – – – – – x x – Undecane – – < 1 – 58 510 – – 7.6 Valeraldehyde – – – – – – 3.1 ± 2.7 0.5 ± 0.3 – Vinylchloride < 0.1 < 0.1 – – – – – – – Xylene 1.2 – 5.7 0.4 – 19 – < 1 66.5 797 x x 10.3 DBP 0.1 – 11.1 < 0.1 – 1.7 – – – – – – – DEHP 0.3 – 1.4 0.2 – 1.1 – – – – – – – Other VOC – – – – 1492 5899 x x 119.6 Other SVOC < 0.001 < 0.001 – – – – – – 9.4

a Wensing (1999) ; new devices. b Funaki et al . ( 2003 ) 20 L chamber, 25 ° C, 50% RH. c Unpublished results (WKI); measurements according to the specifi cations in RAL UZ 122. d Nakagawa et al . ( 2003 ) 1 m 3 glass chamber, n = 1 h − 1 , 22 ° C. e Bak ó - Bir ó et al . ( 2004 )1 m 3 glass chamber, n = 2 h − 1 , 24 ° C, 25%, after 6 h of operation. f Old devices with 2000 h of operation minimum. x: Identifi ed; emission rate not published.


17.3 VOC and SVOC Emissions from Various Devices 417

Table 17.4 Test critera for toner and analytical results (Hahn et al ., 2004 ).

Element/compound Analytical technique

Results ( n = 7) (mg/kg)

Recommended value (mg/kg)

Inorganic Cd ETAAS, ICP - MS 0.361 – < 0.1 5 Co 85 – 9.1 25 Ni 92 – 37 70 Pb 4 – < 1.0 25 Hg AFS, Hydrid - AAS 6.2 – < 5.0 2 Cr ETAAS

UV - VIS (Chromate) 1240 – 19 1

Sn ICP - MS 640 – 1.3 5

Organic Benzene TD – GC – MS 4 – < 1 1 Styrene 159 – < 4 40 Toluene 43 – < 4 40 Ethylbenzene 51 – < 4 40 m - Xylole 46 – < 4 40 a o - Xylole 88 – < 4 p - Xylole 10 – < 4 TVOC > 2100 – 180 1000

a Recommended value for sum of xylole isomers.

the toner requires a complex analytical procedure. Hahn et al . (2004) proposed analytical methods and a list of criteria to evaluate the contents of toner powder (Table 17.4 ). They found BTEX aromatics and styrene in many toner samples. These compounds are also often detected during emission test measurements (Lee, Lam and Fai, 2001 ; Bak ó - Bir ó et al ., 2004 ; Jann et al ., 2005 ). Regarding the heavy metal analysis of toner a wide spread of concentrations could be observed. The exposure against these elements during the printing process would be pos-sible via direct particular emission, but a release of toner dust was not reported for this case (Ewers and Nowak, 2006 ). The risk assessment of direct contact with particulate toner and of the emissions that occur during printing was performed by Roller in 2006. It is known that alveolar and persistent dusts increase the risk of lung cancer independently of their chemical composition (Roller, 2006 ). However, the direct emission of toner dust from laser printers during application could not be observed (Nies, Blome and Br ü ggemann - Prieshoff, 2000 ; Smola, Georg and Hohensee, 2002 ). Test chamber measurements of printers show higher VOC emission during the printing phase than within the idling or feeding phase. For laser printers styrene and xylene are main components whereas alcohols like pentanol were detected to emit from ink - jet printers (Kagi et al ., 2007 ). As men-tioned above, it is known that laser printers can emit aromatic hydrocarbons during operation, of which benzene is the toxicologically most important compo-nent. This carcinogenic compound can be found in headspace analysis of toner


and during the operating phase of laser printers (Jungnickel, Kubina and Fischer, 2003 ). The results of the benzene content in the analyzed toners are spread between 0.1 and 120 mg kg − 1 ( n = 173). The average concentration (3.2 mg kg − 1 ) features the same order of magnitude as reported by Hahn et al . (2004) , see Table 17.4 . A weak correlation between benzene content in the toner and measured air concentration of a printer in the test chamber could be observed but the values in air are higher than expected in all cases. The additional benzene is possibly emitted by other parts of the printer or is formed by reactions during the printing process. The blank - free determination of benzene increases the testing effort of approval measurements because Tenax TA is not a feasible sorbent for this case. Tenax TA shows a small blank of benzene that increases rapidly if a reactive species like ozone is present (Clausen and Wolkoff, 1997 ). For this reason the benzene analysis has to be done separately by an additional sampling with a more stable adsorbent like for example, CarboTrap or CarboGraph (RAL, 2006 ). The estimation of benzene exposure during an 8 h working day was performed by Roller with a long - time average air concentration of 3 μ g/m 3 . Additionally to a theoretical exposure to toner dust in a concentration of 15 μ g/m 3 , the life - time cancer risk would result in 1.1 . 10 − 4 (1 : 9000). As always with carcinogenic compounds, the minimization principle is of priority in such a case. Therefore, in the last few years the manu-facturers have improved toner composition and the printing process to avoid benzene emissions. This effort is rewarded by the environmental label awards (e.g., Blue Angel ). The study by Jungnickel, Kubina and Fischer in 2003 reports benzene emission rates above 0.1 μ g unit − 1 min − 1 (6 μ g unit − 1 h − 1 ) for only 11 laser printers in chamber tests of 65 devices. This demonstrates that many modern laser printers emit benzene only in very low concentrations.

The results of different studies show the necessity of a vast data collective to provide statistical assured conclusions on printer emissions. The main reason is the occurrence of substantial changes in the composition of the consumables. Wensing et al . (2002a) reported chamber test results for 14 laser printers. The TVOC emission rates extended from 76 to 5365 μ g unit − 1 h − 1 (P 90 = 2543 μ g unit − 1 h − 1 ). Only two of the printers would cause a calculated concentration in a model room that exceeds the hygienic precautionary value for TVOC of 0.3 mg m − 3 proposed by Seifert (1999) . The limit requirements of the Blue Angel RAL UZ 122 (2006) , see Table 17.2 are also linked to that TVOC value (Federal Environmental Agency, 2004 ). Another set of emission measurements was published by Jann et al . (2005) and Jann and Wilke (2006) . They analyzed the emissions from 57 different hard - copy devices within the framework of the Blue Angel environmental award. Unfortunately, the raw data of these results were not part of the publication and are therefore unavailable for further statistical analyses. However, the maximum value for each identifi ed substance is easy to determine. A noticeable benzene emission rate of 10 mg unit − 1 h − 1 was detected during the measurement of one singular device. As expected, the TVOC emission rates during the printing process are wide spread between ∼ 0.1 to ∼ 17 000 μ g unit − 1 h − 1 . Beside aromatic hydrocarbons (especially BTEX and styrene) different siloxanes, cyclohexane, and 1 - butanol were identifi ed and quantifi ed. A comparison of the emissions from


laser printers to other electronic devices is shown in Table 17.3 on the basis of the 90 - percentile of 13 different printers. The printers were tested according to the specifi cations in RAL UZ 122. Beside the VOC analytic, dust (0.1 – 1.3 mg h − 1 ) and ozone (0.03 – 2.73 mg h − 1 ) were also measured.

In addition to the emission of toner ingredients the release of organic com-pounds from paper in the printing process is of high interest. Wilke, Jann and Br ö dner (2003) compared laser printer emissions measured in an emission test chamber with the results of direct thermal desorption of paper. Beside the typical woodbased paper emissions (e.g., hexanal and acetic acid) they found high emis-sion rates of SVOC like dibutylphthalate, diisopropylnaphtalene and dioctyladi-pate, especially for recycled paper. The residues of diisopropylnaphtalene, which is typically found in printing inks, result from an incomplete de - inking process of the recycled waste paper. Other SVOC emissions could be linked to chemicals that are used during the clean - up process of the primary paper. From this reason, the use of a defi ned paper is necessary when performing an emission test. In contrast to gaseous emissions, the release of ultra - fi ne particles, which is discussed in the last chapter, is not infl uenced by the type of the paper (Nies, Blome and Br ü ggemann - Prieshoff, 2000 ; Uhde, He and Wensing, 2006 ).

Photocopiers feature some differences compared with desktop laser printers. They are typically constructed for a higher throughput, permanent standby, and immediate print start. Like laser printers the temperature in the device drops when it is in idle - mode and rises fast when a print job arrives. Therefore, the total emis-sion of such a device fl uctuates over a period of time as well.

When comparing the results of different studies about laser printer emission it has to be considered that the infl uence of the consumables on the emitted compounds is very strong. Paper and toner emit various substances under the high temperature conditions during the printing process. Nevertheless, the polymer parts of the laser printer are heated up too. When performing an emis-sion test of such a device it is diffi cult to distinguish between the emissions of the consumables and the hard - copy device. The separate analysis for example, of the toner yields the content of volatile compounds but the inference on the emission during printing is associated with a high uncertainty (Jungnickel et al ., 2003 ). Exactly defi ned testing procedures like published by ECMA (2007) and RAL (RAL, 2006 ) are necessary to get reliable, comparable and reproducible results.

17.3.2 Personal Computers

The emission of chemical compounds from personal computers is of interest because the operating time of such a device in an offi ce or lab may reach > 12 h per day. During operation the CPU inside of a PC can reach temperatures between 80 ° C and 100 ° C with surface temperatures between 50 ° C and 80 ° C (Ren, Cheng and Chen, 2006 ). These temperatures accelerate the release of compounds from plastic materials in the interior of the PC.


The printed circuit boards that are used in personal computers consist of a phenol or epoxide resin base material that contains copper layers. Due to the high operating temperatures these units have to feature a high fi re resistance that is usually obtained by the use of different fl ame retardants. In the United States brominated fl ame retardants are widespread whereas in Europe chlorinated com-pounds are used (Alaee et al ., 2003 ). M ö ller et al . (2004) reported POC emissions from personal computers in the range of 100 ng unit − 1 h − 1 . Compounds with a high percentage of the total emission were TBP, TCEP, TCPP, and TPP. These com-pounds are in use as fl ame retardant and/or plasticizer (Wensing, Uhde and Salthammer, 2005 ). The emissions are presumably caused by the heating of the casing but not by the inner parts. Printed circuit boards are typically equipped with brominated fl ame retardants that are cross - linked with the polymer matrix during fabrication (e.g., Tetrabromobisphenol A). This is advantageous compared with chlorinated compounds because the release of matrix - bounded substances is hin-dered. The emission of brominated and organophosphate fl ame retardants were reported by Kemmlein, Hahn and Jann (2003) . They tested printed circuit boards at a temperature of 60 ° C to simulate operating conditions. The organophosphate emission showed emission rates between 10 and 85 ng unit − 1 h − 1 . Another detected class of compounds were polybrominated diphenylethers like TetraBDE (BDE47) that featured the highest emission rate determined in the study of 14.2 ng unit − 1 h − 1 . During the measurement of these SVOC emissions it has to be considered that emission test chambers may represent a considerable sink to the compounds. Therefore, the authors chose a total testing time of 110 days.

Polymer additives, which are used in high quantities in a plastic material, can contain residues of contaminants from production. Kaiser, Lorenz and Bahadir (1992) found low contents of PAHs within plastic additives. Both classes of com-pounds can be found in dust that accumulates in personal computers. Ren, Cheng and Chen (2006) showed that heated plastic parts are not the only source for PAH because the concentrations of PAH in dust are higher in offi ces were cigarettes are smoked. Anyhow, the study alleged a signifi cantly higher PAH exposure of workers if a PC is present in the room.

Sensory effects of personal computers were investigated by Wargocki et al . (2003) and the impact on IAQ was described by Bak ó - Bir ó et al . (2004) . Chamber test results of the second study are displayed in Table 17.3 . The computers were operated for 3 months under offi ce conditions. The offi ce under observation had a ventilation of 10 l s − 1 and the presence of PCs increased the percentage of dissatis-fi ed test person from 13 to 41%. Additionally, the time for text processing (as a measure for declining work performance) increased by 9%. The compounds with the highest emission rates during chamber tests were phenol and toluene. Both compounds showed model room concentrations below the exposure and odor detection limits. The authors concluded that the impact on the indoor air percep-tion were not a result of the detected chemicals but are caused by so - called ‘ stealth chemicals ’ that were not detected during air analysis. This assumption could not be proved but the results of the SVOC analysis mentioned above show very low concentrations for these compounds that may be not detected during standard VOC analysis.


17.3.3 Television Sets and Computer Monitors

As mentioned above, personal computers and monitors operate over a long period of time during normal offi ce work. It has to be distinguished between two types of monitors: cathode ray tube ( CRT ) monitors and thin - fi lm transistor monitors ( TFT ). Television sets for home usage were usually part of the fi rst category of monitors. Since 2006 the sales fi gures of TFT - based television sets exceed the amount of CRT devices sold in the same year. The emission properties of these new devices are rarely reported until now (e.g., Nakagawa et al ., 2003 ). In 1999 Wensing et al . reported unit specifi c emission rates of 10 CRT - TV sets. New devices showed TVOC values between 189 and 2036 μ g unit − 1 h − 1 and the emission decreases after four months to 25 – 157 μ g unit − 1 h − 1 . Computer CRT - monitors show a similar range in the TVOC values (Table 17.5 ). GC – MS analysis of the obtained chamber air samples identifi ed ∼ 350 VOC and ∼ 250 SVOC compounds. Selected compounds are given in Table 17.1 . Concerning the decay in the emission over

Table 17.5 Emission test chamber measurements of monitors (initial start - up) with sampling after 6 h of operation; SER ( μ g unit − 1 h − 1 ) and model room concentrations ( μ g m − 3 ) (Wensing et al ., 2002a ).

Monitor TVOC Toluene Phenol

SER Model room SER Model room SER Model room

M - 1 526 60 64 7 24 3 M - 2 1 376 158 321 37 130 15 M - 3 1 545 178 600 69 150 17 M - 4 2 231 256 780 90 32 4 M - 5 771 89 231 27 31 4 M - 6 1 025 118 538 62 74 9 M - 7 2 427 279 1045 120 223 26 M - 8 2 534 291 219 25 78 9 M - 9 1 455 167 146 17 48 6 M - 10 1 713 197 758 87 211 24 M - 11 1 136 131 210 24 47 5 M - 12 1 607 185 189 22 20 2 M - 13 a 11 099 1276 207 24 130 15 M - 14 b 253 29 11 1 18 2 M - 15 b 301 35 35 4 59 7 M - 16 b 214 25 32 4 33 4 M - 17 b 36 4 1 1 7 1 M - 18 701 81 147 17 134 15 M - 19 732 84 91 10 35 4

a Main compound limonene. b Old device; total number of hours in operation before testing unknown.


time differences between the VOC and SVOC emissions exist. SVOC emissions like fl ame retardants and plasticisers increase with increasing operating time even if the device is operated longer than one week (Figure 17.5 ). The rising concentra-tions inside the test chamber over an extended period of time clearly prove that these emissions cannot be measured using short - time measuring methods. This rise can be attributed to sink effects inside the test chamber (adsorption on the chamber walls) or to material properties of the studied devices (diffusion from deeper material layers [Wensing, 1999 ]) as well as to a combination of both param-eters. In view of the transferability of the measured SVOC concentrations into the real room, a long - term measurement that allows the emission test chamber to reach equilibrium concentration is required. Compounds with a higher volatility show rapid increase in emission strength after turn - on of the device and then a slow decay over time (Figure 17.6 ). The impact on IAQ can be observed from the calculated concentrations in a model room. These concentrations which are listed in Table 17.5 present an average TVOC value of ∼ 200 μ g m − 3 with a high standard deviation (140%). The assessment of this value was done by Wensing et al . (2002a) on the basis of the hygienic precautionary value published by Seifert (1999) . For the monitors under observation only one exceeded the recommended TVOC value of 0.3 mg m − 3 . Another approach to determine the infl uence of the operation of television sets and monitors in a room was reported by Wargocki et al . (2003) . They examined the sensory impact of four television sets, eight TFT - , and eight CRT - monitors. A personal computer equipped with a TFT monitor had no impact

Figure 17.5 Long - term emission behavior of POC from a TV set (Wensing, Uhde and Salthammer, 2005 ).


even after an operation time of 600 h. The highest sensory emission rates were determined for computers with CRT - monitor (2.7 ± 1.7 olf unit − 1 after 50 h) while the TV sets showed an average value of 1.0 ± 0.6 olf unit − 1 .

Like with other products that contain plastic parts at elevated temperatures, dif-ferent SVOC emissions are observed from monitors and television sets. Typically, phthalate emissions (DBP, DEHP, etc.), chlorinated and non - chlorinated POC can be found (M ö ller et al ., 2004 ; Wensing, 2004 ; Wensing, Uhde and Salthammer, 2005 ), see Table 17.6 . For the evaluation of indoor air pollutants in general a German commission has developed a scheme for calculating indoor air guideline values on the basis of toxicological data (Ad - hoc Working Group, 1996 ). Here, guideline II (RW II) defi nes a value that requires immediate action. Levels below Guideline I (RW I = RW II/10) are considered safe for lifelong exposure. For TCEP, Sagunski and Ro ß kamp (2002) have derived values of RW II = 50 μ g m − 3 and RW I = 5 μ g m − 3 . Under consideration of insuffi cient toxicological data for other POC the authors suggest RW I/II as cumulative value for TCEP, TCPP, TBP, TBEP TEHP and TPP concentrations.

When using the unit - specifi c emission rate of the device with the highest emission (monitor 21, 14 d) in Table 17.6 as a basis for the total of POC (SER U = 2.6 μ g unit − 1 h − 1 ), a theoretical indoor air concentration of 0.3 μ g m − 3 is obtained for the model room (DIN EN ISO 16000 - 9, 2006 ). The indoor air concen-trations of all other devices are below this value by a factor of at least 10. The present results show that the POC emissions then result in a theoretical indoor

Figure 17.6 Long - term emission behavior of VOC from a computer monitor (Wensing et al ., 2002a ).


air concentration which is far below RW I (5 μ g m − 3 ). Carlsson et al . (1997) detected signifi cant increases of TPP concentrations in the indoor air during investigations of indoor air in the presence of new monitors.

A special class of non - reactive additives in polymer materials are CPs, which have plasticizing and fl ame - retarding properties. They were found in an emission test of a television set with a maximum concentration of 2.2 μ g m − 3 after 220 h of operation (Wensing, 2003 ). CP are used as a substitute for PCBs, which have been prohibited worldwide since 2001. The detection of CP in indoor air needs a complex analytical procedure because the amount of single compounds in a CP mixture is high. Therefore, this class is seldom found during standard TD – GC – MS analysis.

The emission of a complete set of personal computers and monitors are described by Nakagawa et al . (2003) . Several VOC like benzene, toluene, etc. were identifi ed and quantifi ed. The results are shown in Table 17.3 . The emission rates of aliphatic hydrocarbons, terpenes, esters, ketones, alcohols and halogens were not found to be signifi cantly different for PCs with CRT and TFT monitors. In the case of aromatic hydrocarbons the emission rates were higher if a PC with CRT monitor was used. The same was found for aldehyde emissions but the differences in emission rates were lower. The separate test CRT monitor and the associated computer in this study proved that the monitor was the main source of chemical emissions.

In the evaluation of these results it has to be considered that the mentioned electronic devices are operated for several hours a day as a set. The sensory impact

Table 17.6 Unit - specifi c emission rates , SER U (ng unit − 1 h − 1 ), for POC of 4 monitors; sampling after different times in operation (7 d, 14 d) in the test chamber (Wensing, 2004 ).

POC a M - 20 SER u

M - 21 SER u

M - 22 SER u

M - 23 SER u

7 d 14 d 7 d 14 d 7 d 14 d 7 d 14 d

TBP 16 16 18 16 12 16 10 12 TEHP < 5 < 5 < 5 < 5 < 5 < 5 < 5 < 5 TBEP < 5 < 5 < 5 < 5 < 5 < 5 < 5 < 5 TCEP < 5 < 5 28 34 < 5 5 < 5 < 5 TCPP < 5 < 5 2318 2465 81 114 21 24 TDCPP < 5 < 5 < 5 < 5 < 5 < 5 < 5 < 5 TPP – – 133 88 76 109 32 23 TCP < 5 < 5 < 5 < 5 < 5 < 5 < 5 < 5 Sum b 16 16 2497 2603 169 244 63 59

a TBP: Tributylphosphate, TEHP: Tris(2 - ethylhexyl)phosphate, TBEP: Tris(2 - butoxyethyl)phosphate, TCP: Tricresylphosphate, TPP: Triphenylphosphate, TCEP: Tris(2 - chloroethyl)phosphate, TCPP: Tris(chloropropyl)phosphate, TDCPP: Tris(dichloropropyl)phosphate.

b Total of TBP, TBEP, TEHP, TPP, TCEP, TCPP, see text.

17.4 Ultra-Fine Particle Emission from Offi ce Devices 425

of an offi ce computer on indoor air is, therefore, a sum of two singular devices. From this reason, it is not useful to measure the emissions of a whole personal computer because it is impossible to differentiate between PC and monitor emission when both are operated in a chamber. The separate emission testing of single electronic devices is also recommended by the ECMA 328 testing standard.

17.4 Ultra - Fine Particle Emission from Offi ce Devices

In addition to the (S)VOC - emission from laser - printers and copiers the release of ultra - fi ne particle s ( UFP ) is also a topic of high interest. Measurements with par-ticle counters show that many laser - printers emit UFP with aerodynamic diame-ters < 0.1 μ m (Wensing et al ., 2008 ). The measurement of UFP with a high time resolution is very important when researching source and deposition mechanism of these particles (Uhde, He and Wensing, 2006 ). A common measuring instru-ment for the detection and classifi cation of UFP used in many studies on this topic is the scanning mobility particle sizer ( SMPS ). This device separates the particles according to their electrical mobility and detects them via laser scattering. The particle size that can pass the classifi er can be changed with time ( ‘ scanning mode ’ ). Consequently, the detector measures only one particle size at a time and the system is limited to the measurement of a slow - changing aerosol (Hinds, 1999 ; Baron and Willeke, 2005 ).

For the measurement of an aerosol undergoing a rapid change in composition like UFPs from a laser - printer several fast techniques are available. These tech-niques base on optical (fast integrated mobility spectrometer, FIMS [Kulkarni and Wang, 2006 ]) or electrical (fast mobility particle sizer, FMPS [Tammet, Mirme and Tamm, 1998 ]) detection. With an electrical low pressure impactor ( ELPI ) parallel real - time detection and impaction for a later chemical analysis is possible but this technique is more commonly used in exhaust analysis (Maricq, Podsiadlik and Chase, 2000 ) than for indoor air. The result of an UFP emission test via FMPS of a laser printer is exemplary shown in Figure 17.7 . The diagram demonstrates a bimodal distribution of the particle size. One mode has an average diameter of ∼ 50 nm whereas the second features particles below 10 nm.

The UFP emission rate of a laser printer cannot be determined using the same techniques as for the (S)VOC case. Here, the superposition of particle release and deposition has to be solved mathematically. Using this method, different types of emitting behavior ( ‘ constant emitter ’ or ‘ initial - burst emitter ’ ) can be identifi ed for the printer under observation (Schripp et al ., 2008 ). It was mentioned above that the determination of the emission rate of a VOC allows the calculation of the air concentration in a model room. For particulate emissions this is not the case. If the deposition in the room is unknown an estimation of the particle concentration development is not possible even though the emission rate of the printer is known (Wensing et al ., 2008 ).


Beside the physical characterization of UFP by particle counting and spect-rometry, the identifi cation of the chemical nature of these particles is also a topic of high interest: On account of the very low mass of these particles this requires a complex sampling and collection procedure in combination with a precise analy-sis, low detection limits and negligible interfering backgrounds. The results from detailed printer studies (Wensing et al ., 2008 ) indicate that the nanometer aerosol is mainly generated in the heated fuser unit of a laser printer. The combination

Figure 17.7 UFP emission profi le of a laser printer, result from FMPS measurement (Wensing et al . 2008 ).

Figure 17.8 GC – MS chromatogram of SVOCs present in the chamber air during printing.

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431

Index

aaccelerated solvent extraction (ASE) 31,

256f.– accuracy 69acetic acid 281ff.acid-base chemistry 318f.acids, organic 281ff., 295, 306, 318, 396f.active sampling 7adhesives– emissions test standards 124– for fl oor coverings 388f.– odor annoyance 177adsorption– interference with analysis 21, 29, 35,

105f.– of organic compounds 21f., 318– surface 21– theory 20f.aerosols 315, 425– from secondary reactions 111, 306ff., 357air exchange rate (ACH) 59, 65, 149, 302f.air fresheners 176, 221, 357f.air handling units (AHUs) 216f., 230f.air pollutants see also volatile organic

compounds (VOCs)– benzene 55ff., 90, 220– classes of 327f.– formaldehyde 54f., 283, 306– inorganic 90, 278ff., 294f.– particulates 21f., 197f., 359– toluene 219f.air quality, indoor (IAQ)– cleaning techniques 66, 91ff., 183, 230f.,

307f.– effect of indoor chemistry 302f., 329– evaluation 189f.– health effects 221, 306, 330ff., 365– in domestic housing 53ff., 349– large chamber testing 111f., 125f., 128


– sensory perception of 165ff., 196, 330ff.– standards 47f., 52, 119f., 207f.air sampling pumps 27aircraft cabins– air quality 165, 367f.– effects of ozone 94ff., 316aldehydes– from carpets 177, 314– from paints 385f.– indoor air levels 206– sampling and analysis 5, 153allergies 335, 386amines– sampling and analysis 154– trimethylamine (TMA) 318f.analysis method development 36ff., 120f.– new product screening 134ff., 182f.– standardization 132ff., 150f., 153, 168artifacts 12– formation 26as low as reasonably achievable (ALARA)

guideline 329, 334ff.asthma 365– incidence 56, 312, 335automobile interiors see car interiors

bbackground contamination 23badge samplers 51f., 58benzene– diurnal changes, outdoor urban 90f.– from laser printers 417f.– measurement of 52, 143biocides– health risks 262f., 288f.– indoor persistence 241f.– sampling 5, 7, 26– sources 260ff., 380, 383, 394– use in preservation 288ff.

432 Index

biomonitoring 261f.Blue Angel environmental label 409, 412f.,

418breakthrough volume 11f., 25building materials– degradation 142, 361, 395f.– emissions testing standards 104, 109f.,

122ff., 394f.– fl oor coverings 388ff.– inorganic products and fi ttings 393f.,

397ff.– life cycle assessment 373, 375f.– paints and coatings 377ff.– VOC emissions from 87ff., 90, 226,

377f.– wood and wood-based products 394ff.buildings– air-conditioning 215f., 231, 278– complaints from occupants 240, 337,

379, 392– diurnal air quality variation 227ff.– new and refurbished 203, 373, 377– sources of VOCs 221, 225ff., 284ff., 367Byne’s disease 282f.

ccalibration 39, 81f., 141f.cancer 262f., 335, 379, 417– known carcinogens 257, 306, 312,

390, 392– risk assessment unit (UR) 208Cannizzaro reaction 284car interiors– conditioning of 149– effect of temperature and driving

conditions 149, 159ff.– emissions test standards 123, 130, 160f.– organochemical sources and effects

147ff., 151, 165– pollutant identifi cation 154f., 159– quantitative emissions measurement

151ff.– test stand conditions 149ff.carbon dioxide (CO2) 197, 215, 291, 333carbon monoxide (CO) 333, 335carbon-based solid sorbents 4, 6carcinogens see cancercardiovascular effects 335, 365carpets 176f., 248, 314ff., 389ff.certifi cation schemes 119, 120, 131, 388– Blue Angel environmental label 409,

412f., 418– ingredients 359, 365f., 368– test protocols 134, 136, 391f.

chambers for laboratory investigations of materials, pollution and air quality (CLIMPAQs) 28f., 111, 168

cleaning of indoor air– monitoring performance 66, 91ff., 233ff.cleaning products 176, 221, 307, 308, 358f.climatic conditions 215f., 239, 276ff.clothes, VOC emissions from 95f., 221,

361, 366f.‘cold spots’ 35cold traps 10, 34, 174f.computers (PCs) 367, 419f.concrete 312, 393f.condensed-phase chemistry 310ff.– hydrolysis 311– oxidation 310confi dence limits (emissions testing) 143conservation strategies (museums)– biocide use 288ff.– climatic control 276ff.– monitoring pollutant levels 292f.– physical containment 275f.continuous fl ow-PAS 76controls 41– charts 41cooking– kitchen surface emissions 316, 397f.– odors 310– toxic emissions from stoves 366cosmetics 362f.cryo-trapping 3f., 12f., 173

ddecipol scale 169ff., 334denuder sampling 26deodorizers see air freshenersdesorption see extractiondesorption electrospray ionization (DESI)

32detection methods– fl ame ionization (FID) 35, 78ff., 365– fl uorescence (FD) 36– mass spectrometry (MS) 7, 32, 35f.,

66ff.– olfactometry 172ff., 367– particle sizers 425– photo-ionization 80f.detergents 176, 359di-(2-ethylhexyl)phthalate (DEHP) 21, 28f.,

252f., 261, 312diffusion, internal 108f., 139f., 158, 384diffusive sampling 7, 47ff.– applications for air quality monitoring

53ff., 204

Index 433

– sampler types and performance 49, 50ff.– theoretical basis 48f.dinitrophenylhydrazine (DNPH) 51dioxins 256f.discrete sampling 74ff.– FTIR/PAS 76f.– nondispersive PAS 74f.disinsection, aircraft 368disposal bins 367dust, surface– chronic reference dose 263, 265– cleaning 196– fi lter sampling 27f., 412– pollutant levels 240ff., 414, 417– sampler 28, 47

eedge effects 140electronic equipment, emissions 180f.

407f., 415f.– computers 367, 419f.– copiers and printers 221, 414, 417ff.,

425ff.– monitors and TVs 421ff.– standards 128f.– testing 112, 180, 408ff.emission mechanism 139emission rate 106f.– area-specifi c 221emission studies– during operation 410f., 422f.– effect of substrate 381ff., 384– life cycle assessment 373, 375f., 399f.– quantifi cation 376f., 411f.– used for exposure limits 388, 392, 394,

409, 413emission test chambers– choice of sampling method 7, 28f., 121– impact of sink effects 105f., 141, 413, 422– models and calculations 106ff., 410ff.– practical applications 109ff., 408ff.– sample preparation and storage 140f.– size classes 101ff., 138f.– with sensory evaluation 168f.endocrine disrupting compounds (EDC)

240f.environmental– test chambers 101ff.– cells 101ff.evaporative emissions 108f., 139, 157f., 384exposure 239, 261ff., 319, 361– indicators 332ff., 336f.– measurement of 7, 58f.extraction, from solid sorbents 8ff., 30f.

fFick’s fi rst law (of diffusion) 48, 392fi eld and laboratory emission cells (FLEC)

28f., 103, 110f.fi lters– choice of, for SVOC/POM sampling 25f.– cleaning of 24– ventilation (HVAC) systems 316fl ame ionization detection (FID) 3, 7, 35,

78ff., 365fl ame retardants 312f., 420fl oor coverings 177, 221, 312– adhesives 388f.– carpets 389ff.– emissions test standards 123, 126, 129– vinyls/polymer 392f.fl uorescence detection (FD) 36fogging– car windscreen 147f., 151, 155, 158f.– in museum displays 287f.– plate 110, 409, 413– precipitate 155formaldehyde– emission from building materials 281ff.,

383, 388, 393ff.– passive badge sampler measurement 54,

58– real-time monitoring 89f.– removal 94formic acid 283f.Fourier Transform IR (FTIR) analysis 76fragrances 304, 357, 365f.

ggas chromatography (GC) 32, 34f., 142f.gas/particle partitioning 20gas-phase organic oxidation chemistry

303f.– hydroxyl radical 308– nitrate radical 309– ozone 303Germany– Indoor Air Quality Guideline 193, 194ff.– IRK/AOLG Ad-hoc working group 189,

194, 197glass wool 9glassware cleaning and storage 24f.glycols/glycol ethers– from paints 379ff., 387f.– sampling and analysis 5, 153guide values– basis for assessment 191f., 263, 293,

328– derivation factors 194f., 197

434 Index

– for total VOC concentration 198ff., 336f., 342f.

– legal status 194ff., 329gypsum-based products 179, 381ff., 394

hhealth risks– indicators of 137f., 332ff., 336f.– of damp buildings 311f.– of pesticides 262f., 288f.– reducing exposure 195ff., 199ff., 230– to children 261, 263f.– toxicology data 192ff., 222, 263f., 328f.– types of 330ff., 365heaters, ceiling 398f.heterogeneous chemistry 313ff.high performance liquid chromatography

(HPLC) 36– with fl uorescence detection (HPLC-FD)

36– with mass spectrometric detection

(LC-MS) 36home (domestic) environments– perception of odors 165ff.– pollutant levels 54ff., 204ff., 240ff.– sources of contamination 260f., 349ff.household products– home care/cleaning 356ff., 363ff.– human activities 351, 354, 356, 366ff.– personal care 361ff., 365f.– product classes 351– usage and emission rate 350– VOC ingredients 352ff.human body emissions 221, 291f., 316humidity– and sample drying methods 13– control, in museums 277f., 291– effect on diffusive sampler performance

49– interference with metal oxide sensors 86hydrocarbons see also polycyclic aromatic

hydrocarbons (PAHs)– cooking stove emissions 366– sampling and analysis 5, 78ff., 83ff.,

152hydrolysis reactions 311f.hydroxyl radicals 308f.hypersensitivity 329, 331

iincenses 363ff.indoor air– chemical reactions 279, 301f., 329– defi nition 190

– exchange rate with outdoor air 59, 149, 159ff., 216, 227

– velocity 49, 109indoor environment– compounds 22– matrices 22indoor material samples 28inorganic atmospheric compounds 278f.inorganic solid sorbents 4, 6insecticides– health risks 262f., 289– ingredients/emissions 356f., 368insulation materials 177ionization potentials 81f.irritation (nose, eye and throat) 330f., 339f.isoprene 69f., 73, 221, 303

jJunge’s equation 20

kkinetics 108f.

llarge volume injection (LVI) 34latex paints 314, 379, 380ff.legal limit values 194lighting, indoor 309limits of detection (LD) and quantifi cation

(LQ) 38f., 143, 182linearity 69linoleum 178f., 310, 392f.linseed oil 176, 178, 310liquid absorption 3f.loss of target compounds 23lowest observed adverse effect level (LOAEL)

192, 194, 263, 328, 341

mmass spectrometry (MS) 32, 35f.– proton transfer reaction (PTR-MS) 7,

66ff.mass transfer models 108f.materials emissions testing 120f.measurement– fi eld monitoring (workplace) 5, 47f., 190,

217, 233– quality standards 207– total (TVOC) emissions 137f., 142, 154,

191– uncertainty factors 49, 52, 69ff., 130,

139ff.medium-density fi breboard (MDF) 394ff.metal oxide (MOX) sensors 83ff.

Index 435

metals, corrosion 279ff., 283method– uncertainty 130– variability 130micro chamber, emissions testing (μ-CTE)

103f.– standards 122, 127, 130multiple chemical sensitivity (MCS) 335museums– environmental factors 273ff.– exhibit preservation strategies 275f.,

288ff.– pollutants in 278ff., 294f., 396f.musk compounds 242f., 365

nneurotoxic effects 335, 365, 379newspapers 351f., 354, 356nicotine 317f., 333nitrate radicals 309f.nitric oxide (NO) 90f.nitrosamines 154no observed adverse effect level (NOAEL)

328, 341

oodorant 165ff.– identifi cation 172– threshold 180odors– analysis 172f., 367– evaluation (sensory testing) 167ff., 310– identifi cation of odorants 175f.– relationship to VOC emissions 165ff.,

180ff., 341– sources 176ff., 314oils– human skin 316– rancidity 310on-column injection (OC) 34organic compounds in outdoor air 90f.organophosphates– in cars (esters) 159ff.– plasticizers/fl ame retardants 243ff.organotin compounds 246f., 414outdoor air quality– and ventilation effectiveness 230, 278f.– diurnal variation 90f.oxidation reactions 310f.oxygen, magneto-acoustic detector

measurement 76f.ozone– cause of VOC sampling artifacts 14f.,

26f., 95

– effect on aircraft air quality 94ff.– emissions from printers 412, 414– initiated chemistry 94f.– reaction with indoor surfaces 111,

313ff.– reactions with terpenes 303ff.– removal from samplers 27

ppaints (and coatings/varnishes) 221, 314– ‘natural’ 386f.– emissions test standards 125, 127, 130– health hazards 377, 379– interior 377– low/zero VOC 387f.– remover 353, 359f.– solvent-based 383ff.– water-based (latex, resin) 379ff.panel (olfactory) testing 167ff.paper 283, 351, 394, 419parallel testing 12, 69, 90particle size, pollutant– effect on inhalation 22– measurement 425particulate organic material (POM)– analysis 32ff.– defi nition 19f., 239– fi lter sampling 23, 27f.– types and sources 22f., 28partition coeffi cient (particle-gas, Kp) 20,

89f.peak concentrations, monitoring 7pentachlorophenol (PCP) 242, 261, 289Perceived Intensity scale (PI) 171perfl uorinated compounds 246, 248personal care products 221– cosmetics 362f.– hygiene 357, 361f.phenolic compounds 240f.photo-acoustic spectroscopy (PAS) 3, 7,

73ff., 87ff.photo-ionization detectors (PID) 3, 80ff.photocatalytic oxidation (PCO) 91ff.photocopiers 221, 419phthalates see also di-(2-ethylhexyl)phthalate

(DEHP)– emission as SVOCs 248ff., 408, 423– sampling and analysis 22, 59, 153plasticizers 248, 312plastics 179f., 420polishes 359ff.polybrominated diphenyl ethers 253f.polychlorinated biphenyls (PCBs) 5, 253,

255f.

436 Index

polychlorinated dioxins/furans 256f.polycyclic aromatic hydrocarbons (PAHs)

22f., 257ff., 420polyphenylene oxides (PPOs)– as solid sorbents 5, 142– odor properties 180, 182polyurethanes (PUs)– as sink/source for VOCs 366f.– carpet underlay 390f.– foams, as solid sorbents 5, 26– odor properties 180POM see particulate organic material (POM)porous polymer sorbents 4ff., 26, 142pressurized liquid extraction (PLE) 30f., 33printing– emissions during operation 410ff., 414,

417ff.– ultra-fi ne particle emission 425ff.proton affi nity 66f.proton transfer reaction mass spectrometry

(PTR-MS) 66ff., 90pumps, air sampling 27purging, of building TVOCs– equipment 230f.– monitoring of effects 233ff.

qquality control systems– products 102f., 148f., 182f.– test methods 37ff., 141quartz wool– for sorbent retention 8f.– in cold traps 10

rradial diffusive samplers 51, 57real-time monitoring– fl ame ionization detection (FID) 78ff.– importance and uses 65f., 87ff.– metal oxide sensors 83ff.– photo-acoustic spectroscopy (PAS) 73ff.,

87ff.– photo-ionization detectors (PID) 80ff.– proton transfer reaction mass

spectrometry (PTR-MS) 66ff., 90reference– health 192– legal 194reference values (background exposure)

192, 203ff., 263regulations (legislation) 119f., 131ff., 143f.,

191, 400– for priority pollutants 334resins 311f., 379, 383f.

respiratory tract– chemical irritants 306, 332– particulates deposition 22Richtwert (RW I and RW II) values

193ff., 423risk assessment– effectiveness 206f., 292f.– toxicology database resources 207f., 263– use of TVOC value 198ff., 333, 336f.rubber materials 177, 389, 393

ssafe sampling volume (SSV) 11f.sampling– analysis equipment 23– artifacts, from reactive gases 14f., 26f., 95– background (blank) determinations 14f.,

23f.– choice of active or passive method 7, 47– fi lter/sorbent methods 8ff., 24ff.– headspace 168, 175, 366, 368– instrumentation 87– objectives 50, 102– orientation 140– sensitivity 3, 32, 174f.– standardization 132ff.– time period 7, 50, 140f.– transport of samples 140, 167– volume control 7, 11f., 27screening (materials emission) 129, 134ff.sealants 177, 367secondary emissions 282ff., 303ff.secondary organic aerosols (SOA) 111,

306ff., 357semi-volatile organic compounds (SVOCs)– analysis 32ff., 412f.– defi nition 19f., 239– emissions testing 28ff., 155– gas/surface partitioning 20ff.– health related properties 22– identifi cation of 155– indoor exposure and health 261ff.– sampling 23ff., 59, 110, 155– types and sources 22f., 158ff., 240ff.,

287f.sensory testing 167ff.Sick Building Syndrome 165, 316, 332sieves– molecular (solid sorbents) 4– stainless steel (in sampler tubes) 9Singapore– building characteristics 216f., 231– compared with EU and US 218ff., 225sink effect 105f., 141

Index 437

solid phase extraction (SPE) membranes 26

solid sorbents– advantages and limitations of use 4, 174f.– artifacts 12– degradation products 14f.– pre-cleaning and storage 13f., 24– sampler design 8– sampling strategies (active/passive) 7– target compound degradation 15– types and properties 4ff., 6, 142– water affi nity 12f.solvent extraction 8, 30f., 159– concentration of extracts 32f., 35, 174f.– from diffusive samplers 51– injection techniques 34– reaction product artifacts 14, 16solvent-assisted fl avor evaporation (SAFE)

175solvent-based paints 379, 383ff.sorbent impregnated fi lters (SIFs) 26specifi c emission rate (SER) 355– calculation 106ff., 223, 411f.– for standard emissions testing 134, 349standards, international– agencies 138– emissions testing methods 8, 104, 122ff.,

409– harmonization 130f., 144– indoor air 47f., 52, 119f., 207f.– quality assurance 207standardization 119ff.stealth chemicals 180, 420supercritical fl uid extraction (SFE) 30f., 291surface adsorption 21surface chemistrysurface dust see dust, surface– heterogeneous (interface) 279ff., 313ff.– in water fi lm 311f.

ttarget compounds 121tarnishing of metals 279ff.temperature– car interior 149, 159ff.– electronic devices 405f., 419f.– museum 278, 292Tenax® 5, 14f., 26, 142terpenes– breakthrough volume determination 11f.– in air fresheners 357f.– in wood resin 177f.– reaction with ozone 111, 176, 303ff.,

317

thermal desorption (TD)– conditioning of equipment 24, 30– coupling with GC analysis 34f.– from diffusive samplers 51f.– in emissions testing 136f.– process, principles of 8ff.thermoplastic polyolefi ns (TPOs) 179tin compounds, organic 246f., 414toluene– as calibration standard 82, 137, 141,

154, 191– concentration in buildings 219f.– removal by PCO 92ff.toner 414, 417ff.total (TVOC) emissions concept 137f., 142,

154, 191– defi nition 374f.– in health risk evaluation 198ff., 333,

336f.– indicator 336toxicity 137f., 191, 328f., 336tropical climate 215f.

uultra-fi ne particles (UFP) 425ff.utilization cycle 190

vvacuum cleaner sampling 27f., 257validation 37ff., 143f.ventilation– effects on air quality 227ff.– evaluation 65, 223– for reducing health risks 196, 203– systems 226, 314, 316very volatile organic compounds (VVOCs)

142, 239, 281, 374volatile organic compounds (VOCs)– as cause of human symptoms 337ff.– defi nition 191, 374– detection techniques compared 3f., 73– emissions levels over time 375– emissions limits/labels 376– emissions test method 133– exposure indoors 329– identifi ed in emissions testing 350ff.– measurement of the sum (ΣVOC)

155f.– semi-quantitative determination 154– sorbent selection 5, 7, 13, 15– source strength 221– stability, during analysis 15f.– types, in indoor air 55, 217ff., 284ff.,

407f.

438 Index

wwater affi nity 12water vapour– effects on materials 277f., 311f.– interference with air sampling 12f.,

78, 151f.water-based paints 379ff.wood preservatives 289wood products– building panels 394ff., 397

– emissions test standards 123, 125f., 129, 394f.

– formaldehyde and organic acid emissions 281f., 311f., 396f.

– odorants 177f.wood-burning stoves 22wool products 354, 366

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