A review of semi-volatile organic compounds (SVOCs) in the ...

A review of semi-volatile organic compounds (SVOCs) in the indoor environment: occurrence in consumer products, indoor air and dustReview
A review of semi-volatile organic compounds (SVOCs) in the indoor environment: occurrence in consumer products, indoor air and dust
Luisa Lucattini a, *, Giulia Poma b, Adrian Covaci b, Jacob de Boer a, Marja H. Lamoree a, Pim E.G. Leonards a
a Department of Environment and Health, VU University Amsterdam, De Boelelaan 1108, Amsterdam, The Netherlands b Toxicological Centre, Department of Pharmaceutical Sciences, University of Antwerp, Universiteitsplein 1, B-2610, Wilrijk, Belgium
h i g h l i g h t s
* Corresponding author. Department of Environme Amsterdam, De Boelelaan, 1108, Amsterdam, The Net
E-mail address: [email protected] (L. Luca
https://doi.org/10.1016/j.chemosphere.2018.02.161 0045-6535/© 2018 The Authors. Published by Elsevier
g r a p h i c a l a b s t r a c t
Information on semi-volatile organic compounds (SVOCs) in consumer products, indoor air and dust was reviewed.
Limited data on concentrations of SVOCs in consumer goods is available, mainly their presence is reported.
The largest obstacle linking SVOCs in products to indoor air/dust ones is the lack on SVOC concentrations in consumer goods.
a r t i c l e i n f o
Article history: Received 22 October 2017 Received in revised form 24 February 2018 Accepted 26 February 2018 Available online 27 February 2018
Handling Editor: R Ebinghaus
Keywords: Consumer products Indoor air Indoor dust SVOCs
a b s t r a c t
As many people spend a large part of their life indoors, the quality of the indoor environment is important. Data on contaminants such as flame retardants, pesticides and plasticizers are available for indoor air and dust but are scarce for consumer products such as computers, televisions, furniture, carpets, etc.
This review presents information on semi-volatile organic compounds (SVOCs) in consumer products in an attempt to link the information available for chemicals in indoor air and dust with their indoor sources. A number of 256 papers were selected and divided among SVOCs found in consumer products (n¼ 57), indoor dust (n¼ 104) and air (n¼ 95). Concentrations of SVOCs in consumer products, indoor dust and air are reported (e.g. PFASs max: 13.9 mg/g in textiles, 5.8 mg/kg in building materials, 121 ng/g in house dust and 6.4 ng/m3 in indoor air). Most of the studies show common aims, such as human exposure and risk assessment. The main micro-environments investigated (houses, offices and schools) reflect the relevance of indoor air quality. Most of the studies show a lack of data on concentrations of chemicals in consumer goods and often only the presence of chemicals is reported. At the moment this is the largest obstacle linking chemicals in products to chemicals detected in indoor air and dust. © 2018 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND
license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
nt and Health, VU University herlands. ttini).
Ltd. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
L. Lucattini et al. / Chemosphere 201 (2018) 466e482 467
1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 467 2. Search criteria . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468 3. Semi-volatile organic compounds in consumer products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468
3.1. Carpets, textiles and clothing . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 468 3.2. Electronics, electrical and electronic components . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 470 3.3. Furniture . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 470 3.4. Building materials and flooring . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 471 3.5. Cosmetics, health care and cleaning products . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 471
4. Semi-volatile organic compounds in indoor dust . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 472 4.1. Phthalate esters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 473 4.2. Synthetic musks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 473 4.3. Polycyclic aromatic hydrocarbons . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 4.4. Polychlorinated biphenyls . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 4.5. Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 4.6. Polyfluorinated alkyl substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 4.7. Brominated flame retardants (PBDEs, EBFRs and other BFRs) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 4.8. Organophosphate flame retardants and plasticizers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 474 4.9. Chlorinated paraffins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 475 4.10. Dechlorane . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 475 4.11. Parabens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 475 4.12. Siloxanes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 475
5. Semi-volatile organic compounds in indoor air . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 475 5.1. Phthalates . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 475 5.2. Synthetic musks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 476 5.3. Polycyclic aromatic hydrocarbons . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 477 5.4. Polychlorinated biphenyls . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 477 5.5. Pesticides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 477 5.6. Polyfluorinated alkyl substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 477 5.7. Brominated flame retardants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 477 5.8. Organophosphate flame retardant and plasticizers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 477 5.9. Chlorinated paraffins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 478 5.10. Siloxanes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 478
6. Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 478 Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 478 Supplementary data . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 478 References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 478
1. Introduction
Indoor air quality (IAQ) can affect everybody's life, and is defined as the quality of the air within buildings and structures. It is important for the health and comfort of building occupants (US EPA, 2014a). The time people spend in both home and the work- ing environment has significantly increased during the past decades (Owen et al., 2010). The number of studies which evaluates the quality of indoor environment has also increased. In the USA, adults spend, on average, 21 h/day indoors whereas children spend, on average, 17e19 h/day indoors (Mercier et al., 2011), conse- quently, the attention to safety has grown in residences (D'Hollander et al., 2010). Over the past 50 years, considerable changes occurred in building materials and consumer products indoors. Many of them are related to the development of new manufacturing lines and the introduction of synthetic polymers and materials that allowed homes and building materials to be made at a reduced cost. At present, plastic items, cleaning products, textiles and electronic devices like computers, televisions, washing machines etc. are commonly found in houses and other indoor places (Weschler, 2009). These consumer goods contain substantial amounts of additives, such as flame-retardants (FRs), plasticizers, antioxidants, and perfluorinated compounds. These chemicals have already been detected abundantly in the indoor environment (D'Hollander et al., 2010).
Exposure to chemicals released from consumer products can occur via inhalation, ingestion or dermal contact (US EPA, 1997).
Intake of contaminated food contributes to a large portion of the overall exposure to environmental pollutants. In addition, indoor air can contain chemicals released from consumer products which can be inhaled. Dermal exposure to chemicals can also occur via direct contact with the skin. Clothing, cosmetics and other personal care products often contain considerable amounts of chemicals which could enter the body through dermal contact (Schettler, 2006). These phenomena largely depend on the structure of the chemicals and their water solubility (Fulekar, 2010).
House dust is a complex mixture of biological material, matter from indoor aerosols and soil particles (US EPA, 1997). Contami- nation of dust can occur via adsorption of chemicals which are present in the air (Schettler, 2006) and via direct contact with consumer products (Butte and Heinzow, 2002; Rauert and Harrad, 2015). For example, the concentration of FRs in dust deposited on electronic equipment was found to be higher than in dust around the same equipment (Brandsma et al., 2013). The mechanism of transfer of DecaBDE to indoor dust was investigated and high concentrations of DecaBDEwere linked toweathering and abrasion of polymers (Webster et al., 2009).
The potential risk of exposure to some compounds in dust may be equal or greater than the exposure via food consumption for toddlers and infants (Hwang et al., 2008).
Indoor dust is a significant sink of semi-volatile organic compounds (SVOCs) which are used in consumer goods. The World Health Organization (WHO) classifies SVOCs as indoor organic pollutants with a boiling point range between 240/260 and 380/
L. Lucattini et al. / Chemosphere 201 (2018) 466e482468
400 C. They differ from volatile organic compounds (VOCs) and very volatile organic compounds (VVOCs) that present boiling point range of 50/100 to 240/260 C and <0 to 50/100 C, respectively (US EPA, 2014b). The US EPA suggests several guidelines to preserve the indoor air quality (US EPA, 2014c, 2015), however to the authors knowledge, a specific indoor air quality index for SVOCs does not exist.
Because of the limited particle dimensions and the high surface area to volume ratio, dust can initially settle on source or non- source surfaces, and consecutively re-suspend in the air (Liu et al., 2016). The transport of indoor air pollutants from sources to settled dust can follow several pathways, such as volatilization from the sourcewith subsequent partitioning to dust (evaporation), abrasion of the product, transferring microscopic fibers or particles to the dust and direct migration by contact between source and dust (Rauert and Harrad, 2015). Simulation tests on the release of SVOCs from consumer products often use test chambers to estimate emissions from materials. In some cases, SVOCs can absorb to the chamber walls due to their low vapor pressure resulting in a lower concentration detected in the air inside the chamber (C. Liu et al., 2013a). Emission of phthalate esters from vinyl flooring was tested in small and large-scale chambers under different temperatures and ventilation conditions. The influence of the temperature and the area/volume ratio on the volatilization of phthalate esters was determined by the gas phase concentration of diisononyl phthalate (DNIP) and bis(2-ethylhexyl) phthalate (DEHP) in two test chambers with different sizes. The large chambers resulted in lower concentrations. SVOCs showed to attach faster to particles than to indoor surfaces (Weschler and Nazaroff, 2008). The rela- tionship between indoor air, dust and surface films, depending on sources, physicochemical properties and indoor environmental characteristics were established by semi-quantitative measure- ments of SVOCs in a test room. SVOC distributions and concentrations were obtained by air, composite dust and furniture surface wipes. Variation on dust concentrations within the room were observed, and spot samples were not necessarily representative for the average room (Melymuk et al., 2016).
One of the first attempts to link organic chemicals in settled house dust to consumer products based on literature studies was a review on exposure assessment (Mercier et al., 2011). In this study, the use, application and source of a selected set of organic chemicals was described.
In the present study, a review of the main classes of SVOCs used in consumer products and present in indoor air and dust is made. The review is an attempt to link information already available for chemicals in indoor air and dust with the source of such chemicals in products. It also highlights the main gaps that hamper this comparison.
2. Search criteria
An initial screening search was performed to identify the main classes of SVOCs present in indoor air, dust and consumer products (i.e. electronics, building materials, textiles, furniture, health care/ personal/cleaning products, cosmetics) using Web of Science database in November 2016. No specific criteria in terms of geographical area or time period were defined (‘general search’). This was followed by a systematic search using as input keywords the classes of compounds followed by the initial screening. We limited our search to studies published during or after the year 2000 in order to capture data that would be most informative on contemporary dust composition. The following classes of com- poundswere included: Polychlorinated biphenyls (PCBs), Polycyclic Aromatic Hydrocarbons (PAHs), Polyfluorinated alkyl substances (PFASs), Polybrominated diphenyl ethers (PBDEs), Brominated
flame retardants (BFRs), Emerging brominated flame retardants (EBFRs), Organophosphate flame retardants (OPFRs), phthalate esters, musks/fragrances, organochlorine pesticides and pyrethroids. The time range of interest was set from January 2000 to November 2016 to have a significant overview of the main SVOCs reported in literature and therefore to extrapolate consistent data related to the same SVOCs present in consumer products. The search terms set in the “title” section were: [PCBs], [PAHs], [PFASs], [PBDEs], [BFRs], [EBFRs], [OPFRs], [phthalate esters], [synthetic musks] or [fragrances], [organochlorine pesticides], [pyrethroids]. The search was extended to indoor air, dust and the selected classes of consumer products as keywords in the “topic” section (e.g. [PCBs] AND [indoor air]). The same criteria were applied to emerging contaminants using as searching terms in the “title” section the following keywords [chlorinated paraffins], [siloxanes], [dechloranes], [parabens]. Literature reviews, modeling papers and test conducted in simulation chambers were excluded. The search results were manually checked and, for the final review, 104 indoor dust and 95 indoor air papers were selected (all the papers can be found in supplementary material). Because of the small number of publications available reporting concentrations of SVOCs in consumer products, the 57 studies reported in the present review are related to the “general search”.
3. Semi-volatile organic compounds in consumer products
The use of SVOCs in building materials, furnishings, electronics, and furniture are often proprietary (usually indicated with the term “additives”), therefore their presence and concentrations is not required to be publically disclosed. This represents a major gap of information. Due to the lack of data related to the concentrations of SVOCs in consumer products, this section refers to the “general search”. For each consumer product category, a brief overview related to the changes in materials and chemicals for the past decades is given, which is relevant for the presence of phased-out or banned chemicals in indoor dust and air.
Table 1 summarized the presence of the selected SVOCs in the classes of investigated consumer products. The green cells indicate the availability of concentrations values, whereas the yellow cells indicate the presence of the SVOC in the consumer product, but no concentration data were reported.
Because of the outcome of the “general search”, the authors decided to separate the “flame retardant class” into PBDEs, emerging BFRs, other BFRs and OPFRs.
3.1. Carpets, textiles and clothing
The carpet industry has changed substantially during the last century. Woven carpets made of cotton and wool have been gradually replaced during the 1950s by tufted carpets made of synthetic fibers such as nylon, rayon and acrylics. By the end of 1960s, the introduction of the olefin carpets led to an increasing use of polyester and polypropylene (Weschler, 2009). In the same period, additives such as fluorinated surfactants as stain repellents and FRs in backing, adhesive and pad started to be introduced in carpets and, since then, their use became common (Weschler, 2009). FRs were also used gradually in fibers, textiles and clothes (Horrocks, 2011; Weil and Levchik, 2008; Weschler, 2009). The evolution of the use of FRs over time is shown in several papers. A gradual replacement of hexabromocyclododecane (HBCD) by the newer FRs, such as OPFRs, was already noticed in 2011 (Kajiwara et al., 2011). Several studies in Japan between 2008 and 2013 revealed the presence of FRs in curtains and textiles, and focused on the mechanistic understanding of possible photodegradation under sun light. High concentrations of HBCD isomers were detected in 9
Table 1 Overview of the applications of SVOCs in consumer products. In green the availability of concentrations values, in yellow availability of information of SVOC in the consumer product with no concentration. White cells indicate no information of the SVOC in the consumer product.
out of 10 tested samples of polyester curtains manufactured in Japan, with concentrations ranging from 22 106 to 43 106 ng/g (Kajiwara et al., 2009). This study suggests the frequent use of HBCD as flame retardant in Japanese textiles. The actual isomeric profiles of HBCD in Japanese curtains were also investigated. a-HBCD was found in higher proportions than in the commercial HBCDmixtures in most of the textile samples. The photolytic transformation of two common FR commercial mixtures (i.e. HBCD and deca- bromodiphenyl ether (DecaBDE)) was tested by exposing polyester curtains to natural sunlight (Kajiwara et al., 2013). The textile samples, purchased from Japanese manufacturers in 2007, showed the stability of HBCD, but also the formation of polybrominated dibenzofurans (PBDFs) suggesting the photodecomposition of DecaBDE under the experimental conditions. The study also showed increasing emission rates of HBCD and DecaBDE from curtains with increasing temperature; noteworthy emissions were detected even at room temperature of 20 C suggesting textiles as potential source of BFRs to dust (Kajiwara et al., 2013).
BFRs in Chinese household products were investigated in 2010 (Chen et al., 2010) showing a high presence of BDE209 (32,611 ng/g) in seat textiles due to the large use of DecaBDE mixture in which BDE209 is the major component. PentaBDE was expected to be present in carpet padding samples. However, its concentration was found to be low, probably due to the lax of flammability standards in China. The high concentration of BDE209 in textile samples was also discussed by Ionas et al. (2015), where levels up to 5.6 105 ng/ g were detected in Belgian carpets and curtains. (Ionas et al., 2015). PBDEs were found in US and German dryer lint samples (median values 803 and 71 ng/g, respectively) identifying dryer electrical components and/or dust deposition onto clothing as potential sources (Schecter et al., 2009). High concentrations of 1,2-bis (pentabromodiphenyl)ethane (DBDPE) and tris(phenyl)phosphate (TPhP) were detected curtains and carpets purchased in Belgium, with maximum values of 2.5 104 and 9.5 104 ng/g respectively. In the same study the presence of 2-ethylhexyl-2,3,4,5- tetrabromobenzoate (EHTBB) was reported with a maximum level of 10 ng/g (Ionas et al., 2015).
Concerning the exact composition of surfactant additives such as per- and polyfluoroalkyl substances (PFASs), in consumer products, this is mostly confidential. During the past years, PFASs in consumer products were studied to enlarge the knowledge on their
content and release (Posner, 2012). In 2009, the US EPA analyzed 116 commercial articles purchased from retail outlets in the United States and grouped them in 13 product categories. Carpets and textiles were classified among the main sources of perfluorocarboxylic Acids (PFCAs) with a maximum concentration of 292 ng/g of fiber carpets and 427 ng/g in product home textile and upholstery respectively (2009 US EPA, 2009a,b). PFASs were detected in 16 jackets produced in Europe and Asia with concentrations ranging from 30 to 458 103 ng/m2 (Lehmphul, 2014). Concentrations and trends of PFASs in carpets, clothing and home textiles were investigated covering a time frame from 2007 to 2011 (Liu et al., 2014). The authors showed a reduction of per- fluorooctanoic acid (PFOA) in each of the product categories analyzed, except for three products (one home textile and upholstery category and two thread-sealant tape products). The presence of PFOA and perfluoroctane sulfonate (PFOS) in different textiles showed the highest concentrations of PFOA and PFOS in the nylon and in cotton samples, with maximum concentrations of 45.9 ng/g and 81.3 ng/g, respectively (Lv et al., 2009). In another study, PFOS in the textiles were higher and between 63 and 13.9 103 ng/g (Lin et al., 2013). Besides PFASs, the textiles also contained pesticides. The presence of the pyrethroid pesticide permethrin on suspended particles indoors has been associated to carpet fiber abrasion (Berger-Preiss et al., 2002). Using permethrin as biocide agent in clothes was verified and confirmed by The Danish EPA in a survey of chemical substances in consumer products of 2014 (Danish Environmental Protection Agency, 2014a). In another survey from the Danish EPA (Danish Environmental Protection Agency, 2006), DEHP was determined in concentrations between 2 103 and 8 103 ng/g in 20 spot samples of textiles of cotton, wool, flax, polyethylene terephthalate (PET) and viscose. The new formula- tions of additives able to confer hydro- and oleofobicity to textile include siloxane (Aslanidou et al., 2016; Lin et al., 2015), while dechlorane plus and chloroparaffin emulsions were listed as suit- able FRagents (Weil and Levchik, 2004, 2008). The use of chlorinated paraffins as additive in textiles was mentioned, but concentrations were not reported (Danish Environmental Protection Agency, 2014b; van Mourik et al., 2016). Additives containing siloxanes for textiles were also reported in literature (Chen et al., 2011; Danish Environmental Protection Agency, 2005).
3.2. Electronics, electrical and electronic components
Electronic device technology has considerably evolved in the past decades. The use of electronics is constantly growing resulting in an increasing emission of chemicals into the indoor environment (Weschler, 2009). Organophosphate and brominated flame retardants have been largely investigated in electronic equipment during the past years.
Kemmlein et al. (2003) (Kemmlein et al., 2003) investigated the emission of FRs from electronic devices simulating an operational condition of 60 C. The authors demonstrated how the emission increases with increasing temperature. Under operational condition, unit specific emission rates (SERu) of organophosphates and polybrominated diphenyl ethers were 10e85 and 0.6e14.2 ng unit1 h1, respectively (Kemmlein et al., 2003).
The presence of PBDEs in Chinese electronics (i.e. television and computer components) and rawmaterial for electronics production was studied by Chen et al. (2010) (Chen et al., 2010). PBDEs derived from the Penta-, Octa-mixture, and Deca-mixtures were detected in 83.3%, 58.3%, and 83.3% of the television casings respectively. Concentrations of PBDEs in computer monitor casings were generally <50 ng/g, except for one sample which contained 13,304 ng/g. The concentrations of PBDEs were higher in computer components (mean value 279,965 ng/g) than television and computer casings, perhaps due to the resistant behavior of plastic materials to high temperatures.
The plastic moldings of TV devices were tested for the presence of BFRs as well as their leaching characteristics in the presence of dissolved humic matter solution (DHM) (humic acid sodium salt dissolved in distilled water and adjusted to 1000mg organic car- bon/liter). (Choi et al., 2009). The PBDE content was about 3% of the total sample weight with DecaBDE being the most abundant ho- mologue (over 80% of the total amount). Tetrabromo bisphenol-A (TBBPA), Polybrominated phenols (PBPs) and Polybrominated biphenyls PBBs were also detected in the same plastic samples in concentrations of 8.1 103, 4.7 103 and 0.25 103 ng/g, respectively. Components of TV sets (e.g. parts of housing front cabinets, rear cabinets and circuit boards) of five sets used in Japan were analyzed in 2008 (Takigami et al., 2008). The highest mean concentrations of PBDEs and TBBPA in the rear cabinets were respectively, 48 106 ng/g and 19 106 ng/g (Takigami et al., 2008).
BFRs and OPFRs were measured in electronic devices of the Japanese market (2008) showing large differences in concentrations ranging from <0.5 to 9.5 106 ng/g for BFRs and <0.9 to 14.0 106 ng/g for PFRs, but also differences in the congener profiles between samples (Kajiwara et al., 2011).
FRs and plasticizers, TBBPA, PBDEs, 1,2-Bis(2,4,6- tribromophenoxy)ethane (BTBPE), tris(phenyl)phosphate (TPhP), Tris (2-chloroisopropyl) phosphate (TCIPP) and tris(methylphenyl) phosphate (TMPP) were also measured with relatively high concentrations (ranging from mg/g to mg/g) in electronic wastes using different analytical methods (Ballesteros-Gomez et al., 2013; Brandsma et al., 2014; Leslie et al., 2016). A pilot study showed trace amounts of PFASs in electronics from Sweden unveiling how PFOS- related substances are still used in a number of applications within the semi-conductor industry and photolithography (e.g. printed circuit boards) (Herzke et al., 2012).
PFOS-based chemicals are often used inmanufacturing of digital cameras, cell phones, printers, scanners, satellite communication systems, radar systems (UNEP, 2007). The presence of PFOS in in- termediate transfer belts of color copiers and printers was reported in concentrations up to 100 ppm. The presence of PFOS in the in- termediate transfer belt suggests that this chemicals are still used by several color copier/multi-function printer manufacturers which dominate the global market and supply spare parts worldwide
(UNEP, 2007). Several studies reported the potential chemicals arising from e-waste disposal recycling. Beside the big concerns related to the toxic metals, levels of organic contaminants are also described. The annual global emission of PCBs in e-waste has been estimated to be 280 tons (assuming a global e-waste production of 20 million tons per year), with the main contribution from recycling of condensers and transformers (PCB concentrations 14 103 ng/g) (Robinson, 2009). Levels of PCBs were determined using a fugacity sampler in an abandoned electronic waste (e- waste) recycling site in South China, with total concentrations of PCBs in the soils of 39.8e940 ng/g, 0.487e8.28 ng/m3 in the air equilibrated with the soil and 0.287e7.38 ng/m3 in the air at 1.5m height from the soil, showing e-waste as a consistent source of PCBs (Wang et al., 2016).
The use of Dechlorane Plus as flame retardant was reported in electronic, wire and cable applications with a content from 5 to 10% (European Union Risk Assessment, 2007). Medium-chain chlorinated paraffins (MCCPs) are used in cable and wire sheathing and insulations as secondary plasticizers in PVC and as softener and FR additives in rubber. The MCCPs used for these purposes usually present an high degree of chlorination (50-42% wt Cl) and are generally added at 10e15% w/w of the total plastic (KEMI, 2017).
Because of their high dielectric constant, siloxanes are considered to be electrically inert (Kamino and Bender, 2013), therefore one of their applications is on electrical materials, but also in sealant coating in domestic appliances such as ovens, irons and refrigerators (Danish Environmental Protection Agency, 2005).
3.3. Furniture
During the last decades, the use of solid wood has been replaced by veneer on composite wood for furniture. Current tables, chairs, desks, dressers, cabinets and bed structures are made of medium density fiberboard or similar composite materials. Synthetic foams treated with FRs are commonly used on cushioning for bedding, sofas and chairs (Weschler, 2009).
Chen et al. (2010) (Chen et al., 2010) assessed the presence of BFRs in sofas, mattresses and pillows. Surprisingly, PBDEs were not found in all polyurethane foam (PUF) samples for furniture and carpet padding, in which PentaBDE mixture was supposed to be used. This was attributed to the lack of recent usage of PBDEs in this type of products due to the restricted furniture flammability standards in China. PBDEs were found in plastic interiors, seat PUF and coating samples collected from cars. The highest value (i.e. 32,611 ng/g) was found for BDE209 in the seat textile due to the large use of DecaBDE in textiles. PBDEs were not detected in all PUF samples for furniture and this disagrees with the assumption that PentaBDE mixture was expected to be used in such products.
In 2012, Stapleton et al. demonstrated how a large volume of new generation FRs was increasingly introduced in US couches, as consequence of the PentaBDE phase out in 2005 (Stapleton et al., 2012). The authors collected and analyzed 102 samples of polyurethane foam from residential couches purchased in the United States between 1985 and 2010. In 41 samples purchased before 2005, 39% of the FRs present were BDEs 47, 99, and 100 (main components of PentaBDE mixture), followed by a 24% of tris(1,3- dichloroisopropyl) phosphate (TDCPP). In the 61 samples purchased in 2005 or later, TDCPP was the most common FR detected (52%) and a mixture of non-halogenated organophosphate FRs, such as TPhP, and tris(4-butylphenyl) phosphate (TBPP) were also found.
Another work conducted by the same author in 2009 reported the content of OPFRs in polyurethane furniture foam in the US (Stapleton et al., 2009). Samples included couches, mattress pads, pillows and chairs; TDCPP was the most abundant compound
(1e5% by weight), followed by tris(1-chloro-2-propyl) phosphate (TCPP; 0.5e22% by weight). Only one sample belonging to the Firemaster 550 flame retardant mixture contained brominated chemicals (4.2% by weight), and one foam sample collected from a futon, likely purchased prior to 2004, contained PentaBDE (0.5% by weight).
A study on the presence of FRs in baby furniture reported the novel FR V6 (2,2-bis(chloromethyl)-propane-1,3-diyltetrakis (2- chloroethyl) bisphosphate) in most of the analyzed samples with a concentration ranging from 24.5 106 to 59.5 106 ng/g of foam (Stapleton et al., 2011).
The use of chlorinated paraffins (CPs) in furniture is restricted by EU regulations (Danish Environmental Protection Agency, 2014b), which suggests the previous use of these chemicals as additives in furnishing materials, but no information is present in the literature beyond the applications in PVC (van Mourik et al., 2016).
3.4. Building materials and flooring
Many existing building materials emit organic contaminants in indoor air. Most of them, such as composite-wood and adhesive resins, were introduced on the market after the World War II. PVC wires and cables insulation were also introduced in the same period replacing rubber and textile braid insulation on wiring and cable. PVC is used in wires and cables of telephone, cable/satellite TV and computer networks systems. Furthermore, plasticizers are added to PVC to make it flexible. To date, PVC pipes (containing organotin compounds as stabilizers) replace copper pipes in drain, waste and water distribution systems (Weschler, 2009).
The content of FRs in building material, mainly insulating foams, was investigated in 2003 (Kemmlein et al., 2003). The study showed that HBCD and TCPP ranged from 1 to 20% in the materials (Kemmlein et al., 2003). Brominated and organophosphate FRs were detected in insulating boards and PVC wallpapers, with the major concentration of HBCDs in the first samples (ranging from 18 106 to 23 106 ng/g) and TPhP in the latters (2.30 102 to 1.8 103 ng/g), except for one wallpaper samples in which HBCD was again the most prevalent FR. Concentrations of all detected FRs were insufficient to impart adequate fire retardancy. PVC on itself is a fire resistant plastic as it has a high chlorine content and therefore FRs are not needed. Therefore, PVC wallpapers seems not to be an important source of indoor pollution by FRs (Kajiwara et al., 2011). On the other hand, short chain chlorinated paraffins (SCCPs) are used as secondary plasticizers and FRs in PVC (US EPA, 2009a,b).
The presence of 1,2-bis(pentabromodiphenyl)ethane (DBDPE), 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE), 2,3- dibromopropyl-2,4,6-tribromophenyl ether (DPTE), 2-ethylhexyl- 2,3,4,5-tetrabromobenzoate (EHTBB), bis(2-ethylhexyl) tetrabromophthalate (BEHTEBP) in XPS construction board, wall paper, parquet underlay, drain pipes, vapor barrier, polyurethane foam and other building material samples was reported (Frederiksen et al., 2014).
The analysis of perfluorooctanoate (PFOA) in consumer articles manufactured with fluoropolymers or fluorotelomer-based products was studied. PFOA was detected only in some samples of mill- treated carpeting indicating that PFOA may not be present or is present only at very low levels (Washburn et al., 2005).
Stone, tile, and wood sealants, and treated floor waxes are important sources of perfluorocarboxylic acids (PFCA), and concentrations in these products range from 4.77 102 to 3.72 103 ng/g product (US EPA, 2009a,b.). PFASs are used in paints, for their water repellent property. Ionic PFASs were found at in two out of the three Norwegian paint samples at low concentrations; the main contribution was from PFOS, with levels of 4.8 and 5.8 ng/g (Herzke et al., 2012). PFASs were also found as
impurities from the production, transport and/or storage (PFHxS and PFBA) in the analyzed paint samples (Herzke et al., 2012).
The emission of DEHP fromvinyl flooring was tested in emission cell studies and showed with a level of ca. 17% (w/w) (Clausen et al., 2012).
A Norwegian study showed the presence of PCBs in building façades, with 2,4-dichlorobiphenyl (PCB7) levels ranging between <1 and 2.9 105 ng/g and <1.94 106 ng/g for plaster and paints, respectively. PCB concentrations varied with the building type and age and were the highest between 1950s and 1960s and decrease in the 1970s. (Andersson et al., 2004). Minor application of dechlorane plus as FR in polyester and epoxy resins (e.g. self-extinguishing phenolic resin laminated paper) was reported (European Union Risk Assessment, 2007). Chlorinated paraffins are used in construction materials such as paints, varnishes, sealants and adhe- sives (e.g. double-glazed windows and dam sealants) (van Mourik et al., 2016). Siloxanes are also used in paints and coatings, conferring UV resistance to the materials (Materne et al., 2005), but also in concrete to reduce the water absorption (Roos et al., 2008), in RTV (Room-Temperature-Vulcanization) silicone sealants (where the siloxanes cover around the 80% of the formulation) (Danish Environmental Protection Agency, 2005).
3.5. Cosmetics, health care and cleaning products
The presence of polyfluorinated compounds (PFCs) was investigated in several household cleaning products, such as impregnating agents (waxes and floor polishes), cleaning agents, lubricants and conditioners. Results showed the presence of at least one PFC in 14 out of 26 samples, with 8: 2 FTOH as dominant PFC (concentrations up to 149mg/mL). PFOA was detected with a maximum concentration of 14.5mg/mL, whereas, surprisingly PFOS was not detected in any sample. PFOA, PFOS, FTOH were mostly found in impregnating agents and lubricants, but were not detected in cleaning agents and conditioners. Impregnating agents containing FTOH showed similar ratios between 6: 2 FTOH, 8: 2 FTOH, and 10: 2 FTOH. FTOH ratios in PFC-containing lubricants were also similar. In 2006, a survey from the Danish EPA (Danish Environmental Protection Agency, 2006) reported the presence of PFCs in several household sprays. PFOSA was detected with concentration of 3.5 103mg/mL in a spray product for impregnation of lather, hide and textiles. The value was in accordance with the declaration of the product (stating that the impregnation product was a fluorocarbon). Another spray product for camping items such as tents and sleeping bags contained PFOS in a concentration of 2.12 101mg/mL. N-ethyl perfluorooctane sulfonamide (EtFOSA) was determined in one out of five liquid floor polishers for vinyl, cork and linoleum with a concentration of 0.1 101mg/mL. Migration tests showed perfluoroheptanoic acid (PFHpA) and PFOA in an impregnation agent mostly used as a liquid in dry-cleaning shops. These two substances have similar properties, similar applications and are released from the same impregnation agent. Eight different PFAS compounds were found in a shoe care agent, reporting low concentrations of PFHpA (1.1 103 ng/g) and PFOA (3.6 103 ng/g).
Phthalate esters DEP (diethyl phthalate) and DBP (dibuthyl phthalate) were detected in body moisture gel and nail gloss samples in percentage between 1.2 and 6.9%. (Chen et al., 2005). Triphenylphosphate (TPhP) was detected with concentrations up to 1.68% by weight in nail polish samples, including two that did not mention TPhP as ingredient (Mendelsohn et al., 2016). Eighteen plasticizers and 12 musks including 10 banned substances were studied in personal care and cosmetics products (Llompart et al., 2013). Twenty-five target compounds (in a total of 30 targets) were found in the samples. The most frequently detected
compounds were two synthetic musks (galaxolide, tonalide, 1.0 106 ng/g) and diethyl phthalate (0.7 103e3.57 105 ng/g). The presence of banned substances (Regulation (EC) No. 1223/2009 and UNION, 2009) such as dibutyl phthalate, diisobutyl phthalate, dimethoxyethyl phthalate, benzylbutyl phthalate, diethylhexyl phthalate, diisopentyl phthalate and dipentyl phthalate, musk ambrette and musk tibetene were detected in sixteen out of the twenty-six personal care products analyzed (62%) (Llompart et al., 2013). Siloxanes are used in cosmetics such as shampoos, conditioners, body/hand/facial creams, deodorants (Danish Environmental Protection Agency, 2005). In 2016, 123 cosmetics and health care products from Portugal were analyzed, and volatile methylsiloxanes (VMSs) were detected in almost all the selected products, with a maximum value of 7.54 105 ng/g in a body moisturized (Capela et al., 2016). Another study from 2011 indicated the linear siloxanes as predominant compounds in a total of 158 personal care products marketed in China, siloxanes were detected in 88% of the samples analyzed, with a maximum concentration of 52.6 106 ng/g in make-up products (Lu et al., 2011a). The use of siloxane as additive in dry cleaning products was also reported (Abelkop et al., 2016). Parabens are widely used in cosmetics as preservative agents (Hu, 2011). The content of parabens in cosmetics showed mean values in the order of mg/L (Wang et al., 2017). Maximum value of 1.65 106 ng/g methyl paraben in a body cream samplewas reported (Melo and Queiroz, 2010). A study from 2008 showed skin care products as the category of cosmetics with the highest parabens concentrations (0.03e0.42% w/w) (Msagati et al., 2008).
Fig. 2. Percentage of studies conducted in different sampling sites.
4. Semi-volatile organic compounds in indoor dust
The distribution of indoor dust samples studied in geographic areas between January 2000 and November 2016 is shown in Fig. 1.
The majority of samples were collected in China followed by USA. Generally, the number of sampling per nation is between 1 and 4. Among the European countries, UK, Belgium and Germany presented a slightly higher sampling rate. On the other hand, areas such as South America and Africa were less studied. This might represent different levels of technology developments between countries or scientific interest, on the other hand these areas are of
Fig. 1. Distribution of indoor dust studies by sampled geographic area. Colors indicate the nu legend, the reader is referred to the Web version of this article.)
interest for the indoor air quality. Concerning sampling site, houses were the main studied sam-
pling sites (45%), followed by offices (12%), shops (7%) and schools (7%) (Fig. 2). A determining factor which contributed to the selection of the sampling site was the relevance for human exposure assessment, covering different targets: families (house dust), adults (office dust) and children (kindergarten and school dust). Houses are the most sampled sites because of the relative easy sample collection protocols. In fact, in many studies, participants were asked to offer dust from their private vacuum cleaner bags for analysis. Concerning shops, the third most investigated sampling site, the site selection is more likely linked to consumer products. The shops covered by our literature search were mainly electronic shops. Electronic shops are becoming a subject of study because a considerable amount of studies focus on FRs and plasticizer emissions.
Themajority of the collected literature (73%) is related to human exposure, biomonitoring, bioaccessibility, and risk assessment
mber of studies per country. (For interpretation of the references to color in this figure
Fig. 4. Ranges of concentrations of the selected SVOCs in indoor dust.
studies indicating the relevance of health risk assessment studies. In Fig. 3 the various microenvironment studies for each SVOC
are shown. The main compounds investigated are FRs, including PBDEs (36%), emerging BFRs (14%), and PFRs (13%). The introduction of new regulations aimed at banning persistent, bioaccumulative and toxic (PBT) chemicals has resulted in a shift of the studied chemicals.While the first studies initially focused on PBDEs (e.g. Rudel et al., 2003; Schecter et al., 2005; Wu et al., 2007), we now observe an increasing interest in PFRs (Canbaz et al., 2016; Hoffman et al., 2015; Van den Eede et al., 2012), emerging BFRs (Ali et al., 2011a; Brommer et al., 2012; F. Xu et al., 2015), dechlorane plus (Cao et al., 2014; W.-L. Li et al., 2015a; Zhu et al., 2007). From 2011 the presence of chlorinated paraffins in indoor dust is increasingly reported (Chen et al., 2016; Friden et al., 2011; Hilger et al., 2013).
A recent systematic literature review and meta-analysis on major SVOCs in the US on indoor dust showed that phthalate esters, phenols, novel FRs, fragrance and PFASs have the highest concentration (Mitro et al., 2016).
The present review shows that SVOC concentrations vary between pg/g and a few mg/g of indoor dust (Fig. 4). In particular, a large variation of ca. six orders of magnitude is observed for BFRs, OCPs, PBDEs, DPs, PCBs, PEs, PFASs and pyrethroids. The classes of SVOCs reaching the highest concentrations are CPs (i.e. SCPs 8.92 105 ng/g, max value) (Hilger et al., 2013), PEs (i.e. SPEs 7.77 106 ng/g, max value) (Rudel et al., 2003), and siloxanes (i.e. Smethylsiloanes 1.16 106 ng/g, max value) (L. Xu et al., 2015), whereas the lowest concentrations were detected for OCPs (i.e. SDDX 5 102 ng/g (Abb et al., 2010), PBDEs (SPBDEs <4 102 ng/g (Van den Eede et al., 2012), PCBs (SPCBs 0.2 ng/g (Abb et al., 2010), PFASs (SPFASs 0.01 ng/g (Eriksson and K€arrman, 2015), pyrethroids (Spyrethroids< 0. 1 ng/g (Rudel et al., 2003), and dechlorane pus (SDPs 0.35 ng/g) (W.-L. Li et al., 2015a).
EBFRs, musks, PAHs and PFRs showed a relatively lower variation within three orders of magnitude, and concentration were in the ng/g range (Abdallah and Covaci, 2014; Fromme et al., 2014a; Kang et al., 2015; Lu et al., 2011b). The large variance observed for SVOCs concentrations in indoor dust is mainly associated with the different sources present indoor. Therefore, this observation highlights the importance that consumer products have in determining the concentration of SVOCs in indoor dust. Beside this reason, the high variation observed in concentrations among studies is related to other three relevant aspects: i) sample type (settle dust, floor dust, dust from vacuum cleaner bags) (Fromme et al., 2014a; Kim et al., 2016; Y. Li et al., 2015b; Shan et al., 2016), ii) sampling method (vacuuming, wipes, brushes, etc.) (Ali et al., 2011a; N. Liu
Fig. 3. Different indoor dust microenvironment studies (%) per selected SVOC.
et al., 2013b; Mannino and Orecchio, 2008) and iii) sample preparation where sieving of dust plays a relevant role in the outcome. Comparing studies is therefore often difficult due to lack of harmonized methods and protocols. Another issue is the expres- sion of sum parameters of classes of SVOCs with different isomers/ congeners between studies. Studies do not always analyze the same sets of compounds from one class. This is a major obstacle when comparing results from different studies.
4.1. Phthalate esters
Phthalate esters were detected at high concentrations in Cana- dian indoor dust. In particular, dibutyl phthalate (DBP), diisoheptyl phthalate (DIHepP), diisononyl phthalate (DINP) and diisodecyl phthalate (DIDP) were detected at maximum levels from 1.32 106
to 1.428 106 ng/g. DEHP was present in a range from 36 103 to 3.84 106 ng/g in 126 dust samples from vacuum cleaner bags (Kubwabo et al., 2013). In the same order of magnitude DEHP was detected in house dust from Kuwait (0.38 106 to 7.8 106 ng/g) (Gevao et al., 2013) and USA (0.17 106 to 7.7 106 ng/g) (Rudel et al., 2003). The same studies from Canada and USA reported relative high concentrations of benzylbutyl phthalate (BzBP) (0.944 and 1.31 106 ng/g, respectively) (Rudel et al., 2003).
4.2. Synthetic musks
Among the synthetic musks, 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8- hexamethylcyclopenta-g-2-benzopyraan (HHCB) is one of the most frequently detected, and generally found at relative high concentrations. HHCB levels up to 11 103 ng/g (Fromme et al., 2004) and 31 103 ng/g (Kubwabo et al., 2012) in German and Chinese house dust (vacuum cleaner bags), respectively were found. The occurrence of synthetic musks in different microenvironments was also investigated (N. Liu et al., 2013b). This study demonstrated a strong link between the source of synthetic musks and indoor dust by showing 10e100 times higher concentrations in barbershop dusts than those from houses, university dormitories and bathhouses: max. concentration levels respectively of the sum of synthetic musks were 1.20 106 ng/g, 1.46 104 ng/g, 6.34 103 ng/g and 4.99 103 ng/g (N. Liu et al., 2013b).
4.3. Polycyclic aromatic hydrocarbons
Polycyclic aromatic hydrocarbons (PAHs) were investigated in indoor dust from various microenvironments such as houses, offices, universities, hospitals, shops and cars, and concentration levels ranging from ng/g to mg/g. For instance, PAHs in floor dust from houses in Guangzhou and Qingyang (China) showed levels from 1.2 to 22 ng/g and 8.5e121 ng/g respectively (Wang et al., 2013c). Higher levels were found in vacuum cleaner dust from Chinese houses (Kang et al., 2015) (14 103 ng/g, median value) and settled dust from Italian houses (Mannino and Orecchio, 2008) with concentrations ranging from 36 103 to 34.453 103 ng/g.
4.4. Polychlorinated biphenyls
The presence of PCBs in house dust was determined worldwide by different sampling methods. The most common samples analyzed are dust from vacuum cleaner bags, “fresh dust” sampled with vacuum cleaner and nylon sampling sock or cellulose extraction thimble (Abb et al., 2010; Abdallah et al., 2013; Rudel et al., 2003, 2008; Takigami et al., 2009; Wang et al., 2013b), followed by settled dust collected with wipes, plastic brushes or dust pans (Tan et al., 2007; Tue et al., 2013; Xing et al., 2011). Low concentrations of PCBs were detected in Singapore, Vietnam and Japan with median values ranging from 5.6 to 23 ng/g (Takigami et al., 2009; Tan et al., 2007; Tue et al., 2013). Higher values up to 13.27 103 and 1.90 105 ng/g were detected in East Germany, West Germany and USA, respectively (Abb et al., 2010; Rudel et al., 2008).
4.5. Pesticides
Chlorinated pesticides and pyrethroids were studied in indoor dust from schools, day cares and homes in North Carolina between 2000 and 2001. High levels of pyrethroids were detected in those samples with a maximum concentration of 3.11 105 ng/g (Morgan et al., 2014). Similar concentrations were found in USA in 2009 (1.72 105 ng/g) (Rudel et al., 2003), while lower levels were detected in another USA study in 2001 (1.5 104 ng/g max value) (Trunnelle et al., 2013). Chlorinated pesticides in dust from North Carolina (4.8 103 ng/g) (Morgan et al., 2014) were comparable to concentrations detected in dust from Danish houses, universities and schools (i.e. 19.03 103 ng/g, max value) (Br€auner et al., 2011). These chlorinated pesticides concentrations were higher than those in house dust from China (i.e. 521 ng/g) (Wang et al., 2013a) and Singapore (i.e. 770 ng/g) (Tan et al., 2007).
4.6. Polyfluorinated alkyl substances
PFASs concentrations in indoor dust from several countries ranged from below the detection limit to 699 ng/g (Eriksson and K€arrman, 2015). A maximum concentration of 121 ng/g was found in Chinese house dust (Shan et al., 2016) and was comparable to levels detected in house dust from the Faroe Islands (i.e. 149 ng/g), Greece (i.e. 129 ng/g), Japan (i.e. 119 ng/g), Spain (i.e. 80 ng/g) (Eriksson and K€arrman, 2015) and Korea (i.e. 97.6 ng/g) (Tian et al., 2016). Among the 17 PFASs studied in Korean house dust, the predominant compounds were PFOS, with concentrations ranging from 0.7 to 52 ng/g, followed by 8:2 FTOH (i.e. 3.1e33 ng/g), EtFOSE (i.e. <LOD-58 ng/g), and PFOA (0.6e11 ng/g) (Tian et al., 2016). Higher PFOS levels were found in house dust from the Czech Re- public (4.8e118 ng/g), where it was also the predominant compound (Karaskova et al., 2016). In the same study, samples from other countries were also investigated and relatively high concentrations of PFHxA were detected in Canadian dust (1.7e146 ng/
g), while PFOAwas the predominant compound in house dust from North America (2.9e318 ng/g) (Karaskova et al., 2016).
4.7. Brominated flame retardants (PBDEs, EBFRs and other BFRs)
PBDEs in indoor dust were subject of many studies (e.g. (Lagalante et al., 2011; Stasinska et al., 2013; Wang et al., 2014). Various countries (Fromme et al., 2014a; Kim et al., 2016; Zhu et al., 2015) andmicro-enviroments (Besis et al., 2014; Kang et al., 2011; Y. Li et al., 2015b) were studied and different sampling method were applied (Imm et al., 2009; Newton et al., 2015; Watkins et al., 2011). The use of vacuum cleaner equipped with a nylon sock (Harrad et al., 2008) is one of the most commonly used sampling methods in the last eight years (Ali et al., 2011b; D'Hollander et al., 2010; Hoffman et al., 2015; Muenhor and Harrad, 2012). This method was applied in Australia, in 2012, to sample schools and houses dust. Median values of 469 ng/g (Toms et al., 2015) and 356 ng/g (Chow et al., 2015) respectively, were detected. Similar results were observed in Poland (323 ng/g, median in house dust) (Krol et al., 2014) and from previous studies conducted in Belgium, where the median concentration levels were 433 ng/g in offices (D'Hollander et al., 2010), and between 313 (D'Hollander et al., 2010) and 360 ng/g (Van den Eede et al., 2011) in homes. Another study conducted in Saudi Arabia showed similar results with median PBDE concentrations of 350 ng/g in house floor dust, 350 ng/g in air conditioning filters and 310 ng/g of dust in cars (Ali et al., 2016).
High concentrations of HBCDwere found in UK house dust, with a maximum of 1250 ng/g corresponding to the a-HBCD isomer (Abdallah et al., 2013). This concentration is similar to maximum values detected in Belgian house dust (1100 ng/g, sampled in 2006 and 1550 ng/g, sampled in 2010) (Van den Eede et al., 2012). Much lower levels were detected in Romanian and Spanish house dust with maximum values of 94 and 34 ng/g, respectively (Van den Eede et al., 2012).
TBBPA was detected in Belgian houses and offices with maximum value of 419 ng/g (D'Hollander et al., 2010); similar concentrations were found in dust samples from two Japanese houses (490 and 520 ng/g) (Takigami et al., 2009).
DBDPE was detected in settled house dust of several cities in Vietnamwithmedian values ranging from40 to 230 ng/g (Tue et al., 2013). DBDPE was also reported in Belgian house and office dust, with median values of 153 and 721 ng/g, respectively; in the same study lower concentrations were found in dust collected from schools in the UK (median 98 ng/g) (Ali et al., 2011a). TBPH (bis(2- ethylhexyl)tetrabromophthalate was detected in house dust from Germany (343 ng/g median value) (Fromme et al., 2014a), while lower concentrations were found in dust sampled in a Thai e-waste storage facility (180 ng/g median value) (Ali et al., 2011b), Belgian houses (13 ng/g median value) and offices (64 ng/g median value) and schools from the UK (96 ng/g) (Ali et al., 2011a).
4.8. Organophosphate flame retardants and plasticizers
OPFRs and plasticizers were studied in indoor dust from several countries. For instance, high concentrations of OPFRs were detected in cars in Saudi Arabia (max 1.09 105 ng/g), while in the same study, lower concentrations were found in house dust (max 1.4 104 ng/g) (Ali et al., 2016). Relatively high levels were also detected in different Spanish micro-environments with PFRs ranging from 2.1 103 to 72.1 103 ng/g (Cristale et al., 2016), which was higher than what was found in Egyptian microenvironments (PFRs 0.96 103e5.24 103 ng/g) (Abdallah and Covaci, 2014).
4.9. Chlorinated paraffins
The presence of chlorinated paraffins was investigated in indoor dust of different microenvironments from Taiwan, showing levels ranging from 1.2 103 to 31.2 103 ng/g (Chen et al., 2016), values comparable to those found in Swedish house dust (i.e. SCPs 3.2 103e18 103 ng/g) (Friden et al., 2011) and in Germany (related to short chain chlorinated paraffins, i.e. SSCPPs 4.0 103e27 103 ng/g), while higher levels were detected for medium chain CPs in German house dust (i.e. medium chain chlorinate paraffins SMCPPs 8.0 103 - 892 103 ng/g) (Hilger et al., 2013).
4.10. Dechlorane
The presence of DP in house dust from China showed a maximum level of 21 103 ng/g (Wang et al., 2011). Value with the same order of magnitude was detected in dust from a Chinese student dormitory (i.e. 14.2 103 ng/g) (Cao et al., 2014). In the same study, lower levels were found in kindergartens (i.e. 231 ng/g) and higher in hotels (i.e. 1.24 105 ng/g) (Cao et al., 2014) (Cao et al., 2014). DP concentrations in Chinese house dust from rural (33e118 ng/g) were similar to the concentrations found in urban areas (2.8e70 ng/g) (Cao et al., 2014). Comparable values were found in house dust from Canada in 2007 (i.e. SDPs 14e61 ng/g), while higher values were detected in dust sampled in 2002e2003 (i.e. SDPs 2.3e5683 ng/g (Zhu et al., 2007).
4.11. Parabens
Relative high concentrations of parabens were detected in indoor dust of micro-environments in Vietnam and house dust from the USA, with Sparaben concentrations between 3.44 103 to 1.06 105 ng/g and 90 to 125.4 103 ng/g respectively (Tran et al., 2016; Wang et al., 2012), while in house dust from China, Korea and Japan lower levels were reported, reaching max values of 26.2 103 ng/g, 11.9 103 ng/g and 19.9 103 ng/g, respectively (Wang et al., 2012). In general methyl-, ethyl-, and propyl-parabens (MeP,EtP, PrP) were found and showing higher concentrations than other paraben isomers (Canosa et al., 2007a, 2007b; Ramírez et al., 2011), with values up to 1.06 105 ng/g, 1.4 103 ng/g and 10.8 103 ng/g for MeP, EtP and PrP, respectively (Fan et al., 2010).
4.12. Siloxanes
The presence of siloxanes was detected in floor dust from different microenvironments (offices, labs, cars) and countries (Tran et al., 2015). The results showed the highest concentration range of total siloxanes (linear and cyclic) (TSi) in Kuwait (i.e. STSi 476e42.8 103 ng/g), followed by Greece (i.e. STSi 340e30.1 103 ng/g). On the other hand, Vietnam and India showed the lowest concentration (STSi nd e 943 ng/g and STSi nd e 657 ng/g, respectively) (Tran et al., 2015). The same study reported levels of total siloxanes in Chinese indoor dust from 117 to 2670 ng/g in indoor (Tran et al., 2015), while a previous study showed a larger concentration range (i.e. STSi 21.5e21 103 ng/g) (Lu et al., 2010).
5. Semi-volatile organic compounds in indoor air
The distribution of the indoor air samples studied over geographic areas is shown in Fig. 5. No studies on dechlorane plus and parabens were found in the literature search, therefore they will not be discussed.
Analogously to the study in indoor dust, the majority of samples
were collected in China and USA (Fig. 5). Generally, the number of studies per nation is between 1 and 3.
Among the European countries, Sweden performed a slightly higher number of studies. As already noticed in the indoor dust paragraph, limited studies were carried out in Africa and South America. Similarly to indoor dust, the microenvironments mostly sampled for indoor air were houses and apartments (49%) and offices (16%) followed by schools and daycare facilities (12%), as shown in Fig. 6.
Themajority of studies were related to human exposure and risk assessment (55%). The main compounds investigated are PBDEs (25%), PAHs (21%), and PCBs (12%) (Fig. 7).
Concentration ranges of SVOCs in all indoor air studies are given in Fig. 8. Generally, SVOCs concentrations vary between several pg/ m3 and a few mg/m3. In particular, a large variation of six orders of magnitude is observed for PAHs and pyrethroids. The classes of SVOCs reaching the highest concentrations are synthetic musks (i.e. Smusks max 3 105 ng/m3) (Lamas et al., 2010), siloxanes (i.e. max 56 103 ng/m3) (Yucuis et al., 2013), PAHs (i.e. SPAHs max 30 103 ng/m3) (Liu et al., 2001) and Pyrethroids (i.e. Spyrethroids max 24 103 ng/m3) (Li et al., 2016), whereas the lowest concentrations were detected for BFRs (i.e. min 0.2 103 ng/m3 (Abdallah and Harrad, 2010), PBDEs (i.e min 0.3 103 ng/m3) (Abdallah and Harrad, 2010), EBFRs (i.e. 2.6 103 ng/m3 min value (Newton et al., 2015), pyrethroids (min 10 103 ng/m3 (Li et al., 2016)) and PCBs (i.e. min 37 103 ng/m3) (Jin et al., 2011).
To assess the SVOC levels in indoor air, easy and reliable sampling methods are necessary (Bohlin et al., 2007). Air samples are usually collected either by active (e.g. high volume samplers - Hi- Vols) or passive samplers (e.g. PUF disks) (Law et al., 2008).
Indoor active sampling methods are accurate and relevant, but they can be intrusive, noisy, and relatively expensive (Bohlin et al., 2008; Tuduri et al., 2012). Passive samplers are cheap, simple to handle, relatively unobtrusive, and a large number can be deployed in different places simultaneously (Bohlin et al., 2008). However, passive samplers sample primarily the gas phase, and are therefore likely to underestimate concentrations of the higher molecular weight SVOCs which are preferentially associated with particulates (Hazrati and Harrad, 2006; Law et al., 2008). In this review, data produced by deploying either active and passive samplers are reported, and no systematic investigation of the potential influence of different sampling methods on the SVOC levels has been conducted.
5.1. Phthalates
High concentrations of PEs were found in indoor hospital air in China (mean total concentration of 19 103 ng/m3) (Wang et al., 2015), 10- to 20-fold higher than those measured in offices (mean concentration of 2.9 103 ng/m3) and apartments (median concentration of 1.1 103 ng/m3) the same geographic area (Song et al., 2015; Zhang et al., 2014). In Western Europe, PE levels were measured at 1 103 ng/m3 in flats and offices from Paris (Moreau- Guigon and Chevreuil, 2014), at 1.1 103 and 1.2 103 ng/m3 in indoor air respectively from apartments and kindergartens in Ber- lin (Fromme et al., 2014b), at 0.68 103 ng/m3 in French residential homes (Dallongeville et al., 2016), at 0.85 103 ng/m3 in indoor air from the University Pierre et Marie Curie in Paris (Braouezec et al., 2016). Similar PE concentrations were observed in indoor air samples collected from Swedish homes (from 1.2 103 to 7.4 103 ng/m3), day care centers (from 1.2 103 to 5.6 103 ng/ m3), and workplaces (from 0.74 103 to 3.9 103 ng/m3) (Bergh et al., 2011), and from the living rooms of eight multi-store apartments in Stockholm, Sweden (up to 2.6 103 ng/m3) (Bergh et al., 2010). Lower PEs levels were finally observed in Japanese offices,
Fig. 5. Distribution of indoor air studies by sampled geographic area. Colors indicate the number of studies per country. (For interpretation of the references to color in this figure legend, the reader is referred to the Web version of this article.)
Fig. 6. Percentage of studies conducted in different sampling sites.
Fig. 7. Different indoor air microenvironment studies (%) per selected SVOC.
Fig. 8. Ranges of concentrations of the selected SVOCs in indoor air.
where indoor air PE maximum concentrations ranged from 0.35 103 to 0.78 103 ng/m3 (Toda et al., 2004).
5.2. Synthetic musks
The occurrence of synthetic musks and fragrance allergens in indoor air was addressed by several studies, all showing the ubiq- uitous presence of this kind of compounds in indoor environments. In Germany, two musk compounds (HHCB and AHTN, acetyl- hexamethyl-tetraline) were detected in the indoor air of kindergartens with median values of 0.10 103 ng/m3 and 44 ng/m3 and maximum concentrations of up to 0.3 103 and 0.11 103 ng/m3
respectively (Fromme et al., 2004). Similar levels of several synthetic musks were measured up to 0.27 103 ng/m3 in primary school classroom in Turkey by (Sofuoglu et al., 2010) and up to 0.11 103 ng/m3 in indoor air from French dwellings (Dallongeville et al., 2016). The presence of HHCB and AHTN was investigated in
indoor air samples collected fromhomes of North-western Spain by (Regueiro et al., 2009), finding concentration ranging from 0.14 103 to 1.13 103 ng/m3 and from 20 to 80 ng/m3, respectively. Both musks were also found in a sample taken in a rest facility from a laboratory building, showing concentration values of 60 ng/m3 for HHCB and 20 ng/m3 for AHTN (Regueiro et al., 2009).
5.3. Polycyclic aromatic hydrocarbons
Low levels of PAHs were measured in French flats, offices, and day nurseries by Moreau-Guigon and Chevreuil (2014), ranging from 0.5 to 1 ng/m3. Higher PAH levels were observed in low energy residential buildings in Lithuania (from 29.7 to 94 ng SPMD/day) (Kaunelien _e et al., 2016), in residential bedrooms in Czech Republic (up to 45 ng/m3) (Melymuk et al., 2016), and in the main living area of houses in Sweden (ranging from 14 to 180 ng/m3) and United Kingdom (from 8.5 to 60 ng/m3) (Bohlin et al., 2008). InMexico City, the levels of PAHs were measured ranging from 12 to 37 ng/m3 and from 6.1 to 92 ng/m3 in indoor air samples collected from the main living rooms of semi-rural and urban residences, respectively (Bohlin et al., 2008). Among the considered studies, the highest PAH levels were measured by (Bohlin et al., 2010) from a Swedish alloy factory (from 320 to 1900 ng/m3) and by (Simcox et al., 2011) in indoor air samples collected from an indoor turf field in Con- necticut, USA (up to 341 ng/m3).
5.4. Polychlorinated biphenyls
PCB levels weremeasured up to 0.14 ng/m3 in indoor air samples of residential bedrooms in Czech Republic (Melymuk et al., 2016) and ranging between 0.35 and 1.8 ng/m3 in air samples collected from French flats, offices, and day nursery (Moreau-Guigon and Chevreuil, 2014). Similar PCB levels were deducted from the analysis of indoor air samples collected from the main living areas of houses located in urban (from 0.210 to 0.840 ng/m3) and in semirural areas (from 0.1 to 0.32 ng/m3) of Mexico City (Bohlin et al., 2008). The levels of PCBs detected in air samples taken from private residences in Gothenburg (up to 1.6 ng/m3) are in the same order of magnitude of those collected from Lancaster (from 0.15 to 2.1 ng/m3) (Bohlin et al., 2008). In France, indoor air samples had concentrations up to 1.5 ng/m3, in line with the other considered studies, showing a widespread distribution of these compounds (Braouezec et al., 2016). Higher PCB levels were finally measured up to 30 ng/m3 in house indoor air from the urban area of Bangkok Metropolitan Region (BMR), while PCB levels comparable with the other considered environments were measured in the suburban (up to 2.1 ng/m3), and rural areas (up to 2.5 ng/m3) of the BMR (Pentamwa and Oanh, 2008). In most of the considered studies, no sources of PCBs were identified in the sampled rooms. Therefore, it is likely that the PCBs detected indoors might have originated outdoors and, when low ventilation existed, the PCBs were accu- mulated because they tend to persist more in the indoor environment (less sunlight).
5.5. Pesticides
The levels of P
11 OCPs were measured in the Bangkok Metro- politan Region (BMR) area by (Pentamwa and Oanh, 2008). In this study, most of analyzed OCPs, except for p,p’-DDD and Mirex, were detected in the BMR urban area (with levels up to 19.4 ng/m3), while lower OCP levels were measured in indoor air samples from rural and suburban homes (up to 0.67 and 2.7 ng/m3, respectively), and many OCP compounds were not detected. In the Czech Re- public, the levels of OCPs found in indoor air sampled from house sleeping rooms reached 0.68 ng/m3 (Melymuk et al., 2016), while in
Lithuanian low energy residential buildings the levels of HCB ranged from 0.7 to 3.1 ng SPMD/day (Kaunelien _e et al., 2016). Comparable OCP levels were measured in air samples from urban residences (ranging from 0.09 to 0.7 ng/m3) and in the living rooms from semirural areas (levels up to 0.36 ng/m3) in Mexico City (Bohlin et al., 2008). Similar OCP levels were measured also in Europe, with concentrations ranging from 0.18 to 0.5 ng/m3 in indoor air samples collected from Swedish houses and between 0.14 and 2.3 ng/m3 in air samples from England (Bohlin et al., 2008).
The presence of pyrethroids was studied in bedrooms air in China with concentration levels ranging from 0.01 ng/m3 up to 24.3 103 ng/m3 (Li et al., 2016), while lower levels were found in air from home and daycare facilities in the USA (i.e. 465 ng/m3 max value) (Morgan et al., 2014).
5.6. Polyfluorinated alkyl substances
The determination of ionic PFAS, including PFOS and PFOA, in the air samples from an office in Hamburg was described by (Jahnke et al., 2007), resulting in maximum PFSA levels of 1.77 ng/ m3. The presence of indoor airborne volatile PFASs, including four fluorinated alcohols (FTOHs), fluorooctane sulfonamides (FOSAs), and fluorooctane sulfonamidoethanols (FOSEs) was investigated in indoor air samples from office environment in Singapore (Wu and Chang, 2012) and PFAS levels were measured up to 6.4 ng/m3.
5.7. Brominated flame retardants
In Tokyo, HBCD was detected at relative high levels, 24 and 29.5 ng/m3, in indoor air samples from house and office environments, respectively (Saito et al., 2007). In the same study, the levels of 2,4,6-tribromophenol (TBPh) (up to 6.8 and 2.8 ng/m3), hex- abromobenzene (HBB) (maximum level of 0.71 and 0.95 ng/m3), and PBDEs (up to 5.9 and 36 ng/m3) weremeasured, respectively, in the same indoor air samples. Higher BFR levels in the office air than the houses were generally observed, likely ascribed to the higher number of emission sources of FRs (fire-resistant interiors, computers, and computer monitors) in the offices than in the houses (Saito et al., 2007). Lower PBDE levels were measured by (Gevao et al., 2006) in house (up to 0.14 ng/m3) and office (up to 0.39 ng/ m3) indoor air samples from Kuwait and by (Braouezec et al., 2016) in indoor environments from the University Pierre et Marie Curie in Paris (maximum level of 0.063 ng/m3). Similar PBDE concentrations were observed in China, up to 0.54 and 0.22 ng/m3 in house and office indoor air samples, respectively (Ding et al., 2016), in the Czech Republic, up to 0.78 ng/m3 in residential bedroom air samples (Melymuk et al., 2016), and from e-waste storage facilities in Thailand (up to 0.35 ng/m3) (Muenhor et al., 2010). From the main living room of several residences, the PBDE levels were comparable and up to 0.46 ng/m3 and 0.20 ng/m3 in urban houses and semirural environment from Mexico City, respectively, up to 0.052 ng/ m3 in Gothenburg, and up to 0.62 ng/m3 from the UK (Bohlin et al., 2008).
5.8. Organophosphate flame retardant and plasticizers
In indoor environments, the PFR levels were usually higher than those of BFRs. Variable levels of PFRs were reported in air samples collected from several indoor environments in Zurich (Hartmann et al., 2004). In this study, total levels of
P 8 PFRs up to 42 ng/m3
were measured in car samples, up to 137 ng/m3 in air samples from a theatre, up to 211 ng/m3 in a furniture store, up to 67.6 ng/m3 in an office building, and up to 91.3 ng/m3 in air samples from an electronic store. These values are similar to those measured in indoor air samples collected from homes (12e240 ng/m3), daycare centers
(14e1.11 103 ng/m3), and workplaces (21e730 ng/m3) from Stockholm (Bergh et al., 2011), and from the living room of multi- story apartments (up to 271 ng/m3) in Sweden (Bergh et al., 2010). Worldwide, comparable levels of PFRs in indoor air samples were measured in Japanese houses and offices, up to 1.507 103 and 260 ng/m3, respectively (Saito et al., 2007), and in Japanese offices from another study (from 124 to 439 ng/m3) (Toda et al., 2004). Slightly higher PFR levels were measured more recently in air samples collected from house living room fromOslo, Norway (up to 1.018 103 ng/m3) (Xu et al., 2016) and in daycare centers from Germany (up to 1.437 103 ng/m3) (Fromme et al., 2014b), showing the increasing role of PFRs in the FRmarket as a replacement for the phased out PBDEs.
5.9. Chlorinated paraffins
Only one study was found that reported the presence of chlorinated paraffins in indoor air from Swedish apartments, with concentrations from <5 to 210 ng/m3 (Friden U.E. et al., 2011).
5.10. Siloxanes
High levels of siloxanes in indoor air of laboratories and offices were found in USA (i.e. 56 103 ng/m3, max value of D5 in one office (Yucuis et al., 2013), while lower concentrations were found in different microenvironments (houses, offices, labs and super- markets) in UK and Italy where the highest values reached 820 103 and 940 103 ng/m3 respectively (Pieri et al., 2013).
6. Conclusions
An overview of the studies conducted on SVOCs in indoor air and dust identified the main literature gaps and the link with chemicals in consumer products.
We selected 104, 95, and 57 studies documenting the presence of selected SVOCs in indoor dust, indoor air, and consumer products, respectively. The identification of the main sources of specific indoor contaminants was not possible because in most cases only the presence was reported, but no concentrations were given. The relevance of the indoor environment quality was highlighted by the common aims, such as human exposure and risk assessment, of most of the studies. This was also reflected by the selection of the sampling sites, with the majority of the studies being houses, offices and schools. Some geographic areas are less represented, such as Africa and South America. Comparison between studies can be hampered due to the lack of harmonized results and protocols (different units, different sample preparation/analytical method, sampling method not standardized), however the lack of data on concentrations of SVOCs in consumer goods still represents the biggest obstacle in linking the sources of chemicals to chemicals in air and dust. The authors propose an interlaboratory study including indoor dust, indoor air and consumer products in order to harmonize sampling protocols and analytical methods and to obtain reliable and harmonized data that could help to fill the gap that currently exists from SVOCs in consumer products and SVOCs in indoor environment.
Acknowledgements
The research for this review has received funding from the Eu- ropean Union's Seventh Framework Programme FP7/2007-2013 under grant agreement no. 316665 (A-TEAM project), and from the European Chemical Industry Council CEFIC through its Long Range Initiative programme, research project LRI-B17, SHINE: Target and non-target screening of chemicals in the indoor
environment for human exposure assessment (2016-2019).
Appendix A. Supplementary data
Supplementary data related to this article can be found at https://doi.org/10.1016/j.chemosphere.2018.02.161.
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