Chemical and Biological Analyses of Selected

Endocrine Disruptors in Wastewater Treatment

Plants in South East Queensland, Australia

Benjamin L. L. Tan B.Sc. (Hons), M.Med.Sc.

Submitted in fulfillment of the requirements for the degree of

Doctor of Philosophy

Australian School of Environmental Studies

Faculty of Environmental Sciences

Griffith University

Queensland

Australia

September 2006

SYNOPSIS

Studies in North America, Europe, Japan and Australia have reported the presence of

endocrine disrupting compounds (EDCs) in wastewater treatment plants (WWTPs) effluent

could affect physiological and reproductive function in exposed fish consistent with

exposure to hormonally active chemicals. The occurrence of EDCs in rivers and receiving

environments situated near WWTPs raises concern over the removal efficacy of these

compounds by conventional treatment processes.

The main aim of this study was to utilize chemical analyses to assess concentrations of

selected endocrine disruptors as well as a biological assay to measure the potential

estrogenic effects of EDCs present in water discharged from wastewater treatment plants in

South East Queensland, Australia. Currently, there are few reported studies on the

estrogenic effects of EDCs released from WWTPs into receiving environments in Australia.

Two field sampling methods were used. Grab sampling with subsequent extraction using a

solid-phase extraction (SPE) technique and passive sampling utilizing EmporeTM (styrene-

divinylbenzene copolymer) disk were used in this study. A gas chromatography-mass

spectrometric (GC-MS) method was successfully developed to simultaneously analyze 15

environmentally ubiquitous EDCs including phthalates, alkylphenols, tamoxifen, androgens

and estrogens. Application of these methods for the determination of target EDCs in

wastewater samples in this study showed 80 99% removal of most EDCs from influent to

effluent, despite the wastewater treatment plants having different treatment processes.

It was observed that the passive samplers accumulated less EDCs than predicted when

compared to the grab samples. This is probably caused by, but may not be limited to,

biofouling, low flow rate, biodegradation and temperature which can progressively reduce

the uptake of compounds into the sampler. A future challenge would be to improve the

reliability of passive samplers by reducing or controlling the environmental conditions that

may impact on the passive sampler performance.

ii

Stir bar sorptive extraction (SBSE) in combination with thermal desorption coupled to GC-

MS was successfully applied to analyze a range of EDCs in wastewater, biosolids and

sludge. The technique was shown to be very versatile, shortening extraction time, reducing

sample volume needed as well as being sensitive for the analysis of a wide range of EDCs.

The results showed that there were high amounts of phthalates, alkylphenols and female

hormones present in the raw influent wastewater and biosolids of the WWTP samples.

For the complimentary bioassay, a proliferation assay using human breast cancer cell line

MCF-7 (E-Screen assay) was used to determine estrogen equivalents (EEqs) in grab and

passive samples from five municipal WWTPs. EEq concentrations derived by E-Screen

assays for the grab samples were between 108 356 ng/L for the influents and

In conclusion, the complementary chemical and biological analyses used in this study

provided a comprehensive assessment which showed that the EDCs discharged from the

monitored WWTPs would be expected to have a low impact on the receiving environments.

Keywords: Wastewater treatment plant; Grab sampling; Passive sampling; Stir bar sorptive

extraction; Gas chromatography-mass spectrometry; E-screen assay; Estrogen equivalent;

Fugacity modelling

iv

ACKNOWLEDGEMENTS It is a great pleasure to thank the many people who made this thesis possible.

It is difficult to overstate my gratitude to my PhD supervisors, Drs. Heather Chapman,

Darryl Hawker, Jochen Mller and Louis Tremblay. With their enthusiasm, inspiration,

patience and great efforts to explain things clearly and simply, they helped to make this

PhD venture a great journey. Throughout my thesis-writing period, they provided

encouragement, sound advice, good teaching, good companies, and lots of great ideas. I

would have been lost without them.

I would like to thank Dr. Frdric Leusch for being a good friend and helping out with most

of the field sampling and laboratory work while at the same time testing the limits of his

olfactory senses. Many thanks go out to Rene Diocares for technical advice and support on

the GC-MS, Katherine Trought, Tamara Ivastinovic, and Ngari Teakle for their help in the

E-Screen assay. I am grateful to the faculty and staff at Griffith University, National

Research Centre for Environmental Toxicology (EnTox), Landcare Research, NZ and

Queensland Health Pathology and Scientific Services (QHPSS) who have made this project

enjoyable, especially Eri Takahashi, Dr. Aedah Abu Bakar, Henrique Anselmo, Brad

Polkinghorne, Eva Holt, Heather Brown, Jason Dunlop, Mary Hodge, Anita Kapernick,

Scott Stephens, Andrew Watkinson, Dr. Simon Costanzo and Colm Cahill. Special thanks

also go to my lively aikido mates for helping me keep calm and collected during the testing

times of my PhD, especially Dr. Daniel James, Steve Dows, Dr. Bruce Tranter, Dan Brown,

Gary Weigh, Gabrielle Paynter, Chris Cobban and Tim Piatkowski.

I would like to acknowledge with gratitude Griffith University, Corporative Research

Centre for Water Quality and Treatment (CRC WQT), EnTox, QHPSS and the Australian

Research Council (ARC) for their financial support and for giving me a chance to

contribute towards the field of Environmental Toxicology.

Lastly, and most importantly, I wish to thank my parents, Susan and David, and sisters,

Yvonne and Yvette, for their love, guidance, support, encouragement and patience

throughout my life. To them I dedicate this thesis.

v

DECLARATION OF ORIGINALITY

The experimentation, analyses, presentation and interpretation of results presented in this

thesis represent my original work that has not previously been submitted for a degree or

diploma in any university. To my best knowledge and belief, this thesis contains no

material previously published or written by another person except where due reference is

made within the thesis itself.

______________________________________________

(Benjamin L.L. Tan)

vi

TABLE OF CONTENTS

Synopsis ii

Acknowledgement v

Declaration of originality vi

Table of contents vii

List of tables xii

List of figures xiv

List of abbreviations xviii

Publications resulting from this research xxi

Other publications related to this research xxii

Chapter 1: Thesis objectives 1

1.1 General introduction 1

1.2 Aims and objectives 3

1.3 Research questions 4

1.4 Thesis format 5

1.5 References 6

Chapter 2: Literature review 8

2.1 Introduction 8

2.2 The endocrine system 12

2.3 Endocrine disrupting compounds (EDCs) 13

2.4 Chemical properties of selected endocrine disruptors 14

2.4.1 Estrogens 19

2.4.2 Tamoxifen 21

2.4.3 Androgens 22

2.4.4 Alkylphenols 23

2.4.5 Phthalates 25

2.5 Endocrine disruption 26

2.5.1 Mechanisms of endocrine disruption 26

2.5.2 Other factors affecting the activity of endocrine disruption 28

vii

2.5.3 Endocrine disruptors in wildlife (vertebrates/invertebrates) 29

2.5.4 Endocrine disruptors in discharge and surface water 31

2.6 Methodologies for detection and monitoring of endocrine disruptors 33

2.6.1 Chemical analytical techniques 33

2.6.1.1 Extraction methods for water 33

2.6.1.1.1 Direct sampling: solid phase extraction (SPE) 34

2.6.1.1.2 Passive sampling 35

2.6.1.1.3 Stir bar sorptive extraction (SBSE) 38

2.6.1.2 Extraction methods for sludge 38

2.6.1.3 Gas chromatography-mass spectrometry (GC-MS) 39

2.6.2 Biological testing 40

2.6.2.1 In vitro bioassays 40

2.6.2.1.1 Receptor binding assay 40

2.6.2.1.2 Estrogen receptor (ER) activation assays 41

2.6.2.2 Whole animal assays (in vivo) 42

2.7 EDCs fate modelling 43

2.8 Risk assessment of EDCs 44

2.9 Conclusions 46

2.10 References 46

Chapter 3: Evaluation of grab and passive sampling methods to determinate selected

endocrine disrupting compounds in municipal wastewaters 66 3.1 Abstract 66

3.2 Introduction 66

3.2.1 Sampling kinetics of EDCs with the EmporeTM disk sampler 69

3.3 Materials and methods 71

3.3.1 Chemicals and reagents 71

3.3.2 Sample collection 72

3.3.3 Processing of grab samples 73

3.3.3.1 SPE extraction procedure 73

3.3.3.2 Grab sampling SPE recovery experiment 74

3.3.4 Processing of passive samples 75

3.3.4.1 Passive sampler pre-deployment conditioning 75

viii

3.3.4.2 Passive sampler calibration experiment 75

3.3.4.3 Passive sampler extraction 76

3.3.5 Derivatization procedure 76

3.3.6 GC-MS analysis 77

3.4 Results and discussion 79

3.4.1 Calibration of passive sampler 79

3.4.2 Environmental monitoring 87

3.5 Conclusions 96

3.6 References 97

Chapter 4: Stir bar sorptive extraction and trace analysis of selected endocrine

disrupting compounds in water, solids and sludge samples by thermal desorption with

gas chromatography-mass spectrometry 103

4.1 Abstract 103

4.2 Introduction 103

4.3 Materials and methods 105

4.3.1 Chemicals and reagents 105

4.3.2 Instrumentation 105

4.3.3 SBSE procedure 109

4.3.4 Sludge/water partitioning experiment 110

4.3.5 Environmental monitoring 111

4.4 Results and discussion 111

4.4.1 SBSE recovery and partitioning experiments 111

4.4.2 Environmental monitoring 113

4.5 Conclusions 117

4.6 References 117

Chapter 5: Comprehensive study of selected endocrine disrupting compounds using

grab and passive sampling at selected wastewater treatment plants in South East

Queensland, Australia. 1. Chemical analysis 121

5.1 Abstract 121

5.2 Introduction 121

5.3 Materials and methods 124

ix

5.3.1 Chemicals and reagents 124

5.3.2 Sampling sites 124

5.3.3 Grab sample collection and extraction 125

5.3.4 Passive sampler conditioning and extraction 128

5.3.5 GC-MS derivatization procedure 129

5.3.6 GC-MS analysis 129

5.3.7 Centrifuged solids and sludge analysis 130

5.4 Results and discussion 132

5.4.1 Grab sampling 136

5.4.2 Passive sampling 141

5.4.3 Solids and sludge analysis 145

5.5 Conclusions 147

5.6 References 147

Chapter 6: Comprehensive study of selected endocrine disrupting compounds using

grab and passive sampling at selected wastewater treatment plants in South East

Queensland, Australia. 2. In vitro biological screening 153

6.1 Abstract 153

6.2 Introduction 153

6.3 Materials and methods 155

6.3.1 Sampling sites 155

6.3.2 Grab sample collection and extraction 156

6.3.3 Passive sampler conditioning and extraction 157

6.3.4 Cell proliferation assay 158

6.4 Results and discussion 161

6.4.1 Estrogenic activity of WWTPs samples 161

6.4.2 Comparison between the estrogenic activity of passive sampler and grab

sampler 165

6.4.3 Comparison of E-Screen assay and analytical chemistry 166

6.5 Conclusions 170

6.6 References 171

x

Chapter 7: Modelling of the fate of selected alkylphenols and phthalates in a

municipal wastewater treatment plant in South East Queensland, Australia 176

7.1 Abstract 176

7.2 Introduction 176

7.3 Process description 180

7.4 The fugacity approach 184

7.5 Results and discussion 189

7.6 Conclusions 197

7.7 References 198

Chapter 8: General discussion and conclusion 201

8.1 General discussion 201

8.2 General conclusion 205

8.3 Future research 206

8.4 References 207

xi

LIST OF TABLES

Table 2.1. Biochemical properties of selected endocrine disruptors 16

Table 2.2. Daily excretion (g) of estrogenic steroids by humans 20

Table 3.1. Retention time and ions used for quantification in GC-MS detection of the

selected EDCs and their respective recoveries with SPE and EmporeTM disk extractions 78

Table 3.2. Selected physiochemical properties and sampling rates of test analytes for the

passive sampler (EmporeTM disk) based on the laboratory calibration at 24C 84

Table 3.3. Concentration of EDCs detected in grab samples from WWTP J 92

Table 3.4. Concentration of EDCs detected grab samples from WWTP M which practices

water recycling 93

Table 3.5. Concentration of EDCs detected using grab and passive sampling methods in the

wetlands of WWTP N 94

Table 4.1. Log Kow, theoretical recoveries, spiked sludge recoveries, retention time, ions

used for quantification in SBSE GC-MS detection 107

Table 4.2. EDCs concentration present in raw influent, anaerobic, aerobic and anoxic zones

of the bioreactor at WWTP J determined by SBSE 116

Table 5.1. Description of the 5 activated sludge wastewater treatment plants in this

study 125

xii

Table 5.2. Log Kow, retention time and ions used for quantification in GC-MS analysis for

the detection of the selected EDCs and their respective extracted recoveries with SPE and

EmporeTM disk, and sampling rates for the target compounds using the EmporeTM

disk 127

Table 5.3. Log Kow, theoretical recoveries, retention time, ions used for quantification in

SBSE GC-MS analysis and detection 132

Table 5.4. Selected analytes present in WWTP A 133

Table 5.5. Selected analytes present in WWTP B 134

Table 5.6. Selected analytes present in WWTPs C, D and E 135

Table 6.1. Description of the 5 conventional activated sludge wastewater treatment plants in

this study 156

Table 6.2. Estrogenicity of individual compounds when tested with E-Screen assay 160

Table 6.3. Aqueous estrogen equivalent comparison between the chemical and biological

analyses and the different sampling methods 164

Table 7.1. Measured phthalates and alkylphenols present in the WWTP A 182

Table 7.2. Selected physical properties at 25 C of the phthalates and alkylphenols used in

this study 183

Table 7.3. Estimated and measured removal efficiencies of selected compounds with

endocrine disrupting properties in WWTP A 192

xiii

LIST OF FIGURES

Figure 2.1. Water supply catchments for nearly half of South East Queensland, Australia

region (SEQ Water, 2002) 11

Figure 2.2. Water storage status of Wivenhoe, Sometset and North Pine Dams suplying

potable water to South East Queensland, Australia (SEQ Water, 2006) 12

Figure 2.3. Chemical structure of natural estrogens 17

Figure 2.4. Chemical structure of androgens 17

Figure 2.5. Chemical structure of pharmaceutical drugs 18

Figure 2.6. Chemical structure of alkylphenols 18

Figure 2.7. Chemical structure of phthalates 18

Figure 3.1. Chromatogram of the selected EDCs 79

Figure 3.2. Examples of time series uptake data in SDB-RPS EmporeTM disk for selected

EDCs in calibration experiment (a) nonylphenol; (b) bisphenol A; (c) estrone and (d)

dibutyl phthalate 85

Figure 3.3. Relationship between log Kow and log KSW for SDB-RPS EmporeTM disk,

including available literature data (Verhaar et al., 1995; Green and Abraham, 2000; Mayer,

2000; Stephens et al., 2005) 86

Figure 3.4. Correlation between measured EDC concentration obtained from grab sampling

and passive sampling at different sites along WWTPs A, B, C, D and E in South East

Queensland, Australia 87

xiv

Figure 3.5. Examples of variation of concentrations for (a) 4-tert-octylphenol and (b)

nonylphenol using grab sampling at WWTP M at selected time intervals over a period of 7

days in 2005 95

Figure 4.1. A. SBSE desorption chromatogram of phthalates, tamoxifen, acyl derivative of

alkylphenols, estrogens and androgen. B. Total ion chromatogram (A) enlarged to show

smaller compound peaks of the chromatogram (the chromatogram for tamoxifen was

removed to give a clearer view of androsterone and etiocholanolone) 108

Figure 4.2. Recovery of EDCs in water and sludge phases when 1 L of bioreactor sample

from WWTP J was spiked at a concentration of 500 ng/L (mean standard deviation) 112

Figure 4.3. Changes in the log Kp (sludge/water partition coefficient) of EDCs onto

bioreactor sludge with log Kow (octanol/water partition coefficient) 113

Figure 5.1. Elimination of estrogens during passage through the 5 WWTPs located in

Southeast Queensland, Australia. -E2 = 17-estradiol, -E2 = 17-estradiol, BDL = below

detection limit. Grab samples collected were from the influent (Inf), bioreactor-anaerobic

(bio-ana), bioreactor-aerobic (Bio-ae), bioreactor (Bio), return activated sludge (RAS),

clarifier (Clar), effluent (Eff), point of discharge in the river or outflow (Dis) and 1 km

downstream from outlet (Riv) 141

Figure 5.2. Correlation between measured EDCs obtained from grab sampling and passive

sampling at different sites at WWTPs A, B, C, D and E 145

Figure 6.1. Comparison of the estrogen equivalent concentration (EEq) determined in the

E-Screen assay with those calculated from the results of chemical analysis of the grab

samples from the influent and effluent of selected five WWTPs in Southeast Queensland,

Australia (Chapter 5). Columns represent the mean standard deviation. Inf = influent, Eff

= effluent, BDL = below detection limit 165

xv

Figure 6.2. Correlation between the measured E-Screen assay estrogen equivalent

concentrations (EEq) and the predicted EEq from the results of the grab and passive

samples from all five WWTPs 169

Figure 6.3. Contribution of steroidal estrogens to total estradiol equivalent concentration

(EEq) calculated from results of GC-MS in selected WWTP samples of five WWTPs in

Southeast Queensland, Australia. -E2 = 17-estradiol, -E2 = 17-estradiol. Grab samples

collected were from the influent (Inf), bioreactor-anaerobic (bio-ana), bioreactor-aerobic

(Bio-ae), bioreactor (Bio), return activated sludge (RAS), clarifier (Clar), effluent (Eff),

point of discharge in the river or outlet (Dis) and 1km downstream from outlet (Riv) 170

Figure 7.1. Diagram of (A) water (m3 h-1) and (B) solids (g h-1) balances for WWTP A.

65% biosolids removal in the primary settling tank is assumed 183

Figure 7.2. Diagram of fugacity transport/process parameters (D) in WWTP A. P =

primary settling tank, Bio = bioreactor, F = final settling tank, B = biodegradation, V =

volatilization 188

Figure 7.3. Process details of fate, D (mol Pa-1 h-1), f (Pa), k (h-1) and Z (mol m-3 Pa-1), for

(A) diethyl phthalate and (B) dibutyl phthalate in WWTP A. Data in bold are the fluxes for

the various processes (g h-1). ZW of diethyl phthalate and dibutyl pthalate are 37.2 mol m-3

Pa-1 and 11.16 mol m-3 Pa-1, respectively 193

Figure 7.4. Process details of fate, D (mol Pa-1 h-1), f (Pa), k (h-1) and Z (mol m-3 Pa-1), for

(A) benzyl butyl phthalate and (B) di-(2-ethylhexyl) phthalate in WWTP A. Data in bold

are the fluxes for the various processes (g h-1). ZW of benzyl butyl phthalate and di-(2-

ethylhexyl) phthalate are 13.0 mol m-3 Pa-1 and 0.576 mol m-3 Pa-1, respectively 194

Figure 7.5. Process details of fate, D (mol Pa-1 h-1), f (Pa), k (h-1) and Z (mol m-3 Pa-1), for

(A) nonylphenol and (B) 4-tert-octylphenol in WWTP A. Data in bold are the fluxes for the

various processes (g h-1). ZW of nonylphenol and 4-tert-octylphenol are 9.09 10-2 mol m-3

Pa-1 and 1.80 mol m-3 Pa-1, respectively 195

xvi

Figure 7.6. Process details of fate, D (mol Pa-1 h-1), f (Pa), k (h-1) and Z (mol m-3 Pa-1), for

(A) 4-cumylphenol and (B) bisphenol A in WWTP A. Data in bold are the fluxes for the

various processes (g h-1). ZW of 4-cumylphenol and bisphenol A are 5.03102 mol m-3 Pa-1

and 1.74105 mol m-3 Pa-1 196

Figure 7.7. Correlation between the estimated and measured effluent compound

concentrations from WWTP A 197

xvii

LIST OF ABBREVIATIONS

Andr. = Androsterone

APE = Alkylphenol ethoxylate

BAC = Biologically activated carbon

BBP = Benzyl butyl phthalate

BDL = Below detection limit

BNR = Biological nutrient removal

BPA = Bisphenol A

BSTFA = N,O-bis-(trimethylsilyl)trifluoroacetamide

CD-FBS = Charcoal-dextran treated fetal bovine serum

CP = 4-Cumylphenol

CV = Coefficients of variation

DBP = Dibutyl phthalate

DDT = Dichlorodiphenyltrichloroethane

DEHP = Di-(2-ethylhexyl) phthalate

DEP = Diethyl phthalate

DNA = Deoxyribonucleic acid

DOP = Dioctyl phthalate

E1 = Estrone

E2 = 17-Estradiol

E3 = Estriol

EC50 = Effective concentration which produces 50% of the maximum possible response

EDC = Endocrine disrupting compound

EC95 = Effective concentration which produces 95% of the maximum possible response

EE2 = 17-ethynylestradiol

EEq = Estrogen equivalent

ER = Estrogen receptor

ERBA = Estrogen-receptor binding assay

Etio. = Etiocholan-3-ol-17-one

FST = Final settling tank

GAC = Granular activated carbon

xviii

GC-MS = Gas chromatography-mass spectrometry

GPC = Gel permeation chromatography

HEPES = 4-(2-hydroxyethyl)-1-piperazineethanesulfonic acid

HPLC = High performance liquid-chromatography

HS-SBSE = Headspace stir-bar sorptive extraction

i.d. = Internal diameter

IC50 = The concentration required to inhibit 17-estradiol binding by 50%

LLE = Liquid-liquid extraction

LOEC = Lowest observed effect concentration

Log Kow = Log octanol/water partition coefficient

NA = Not analyzed

NOAEL = No observed adverse effect level

NOEC = No observed effect concentration

NP = Nonylphenol

OP = 4-tert-Octylphenol

PAH = Polyaromatic hydrocarbon

PCB = Polychlorinated biphenyls

PE = People equivalent

PEC = Predicted environmental concentration

PNEC = Predicted no effect concentration

POCIS = Polar organic chemical integrative sampler

PRC = Performance reference compound

PST = Primary settling tank

RAS = Return activated sludge

RPE = Relative proliferative effect

rpm = Revolutions per minute

RPP = Relative proliferative potency

SBSE = Stir-bar sorptive extraction

SD = Standard deviation

SIM = Selected ion monitoring

SPE = Solid-phase extraction

SPMD = Semi-permeable membrane device

SPME = Solid-phase microextraction

xix

TIE = Toxicity identification evaluation

TMS = Trimethylsilyl

UK = United Kingdom

USA = United States of America

UV = Ultraviolet

VOC = Volatile organic chemicals

VTG = Vitellogenin

WWTP = Wastewater treatment plant

xx

PUBLICATIONS RESULTING FROM THIS RESEARCH Tan, B.L.L., Hawker, D.W., Mller, J.F., Leusch, F.D.L., Stephens, B.S., Tremblay, L.A.,

Chapman, H.F., (submitted for publication). Evaluation of grab and passive sampling

methods to determinate endocrine disrupting compounds in municipal wastewaters.

Tan, B.L.L., Hawker, D.W., Mller, J.F., L.A., Chapman, H.F., (submitted for publication).

Stir bar sorptive extraction and trace analysis of selected endocrine disruptors in water,

biosolids and sludge samples by thermal desorption with gas chromatography-mass

spectrometry.

Tan, B.L.L., Hawker, D.W., Mller, J.F., Leusch, F.D.L., Tremblay, L.A., Chapman, H.F.,

2007. Comprehensive study of endocrine disrupting compounds using grab and passive

sampling at selected wastewater treatment plants in South East Queensland, Australia.

Environ. Int. doi:10.1016/j.envint.2007.01008.

Tan, B.L.L., Hawker, D.W., Mller, J.F., Leusch, F.D.L., Tremblay, L.A., Chapman, H.F.,

2007. Modelling of the fate of selected endocrine disruptors in a municipal wastewater

treatment plant in South East Queensland, Australia. Chemosphere

doi:10.1016/j.chemosphere.2007.02.057.

xxi

OTHER PUBLICATIONS RELATED TO THIS RESEARCH

Leusch, F.D.L., Tan, B.L.L., Tremblay, L.A., Chapman, H.F., 2005. Endocrine disruptors

in sewage: Perception vs. reality. Proceedings of AWA OzWater 2005, 8-12 May 2005,

Brisbane, QLD, Australia.

Tan, B.L.L., Hawker, D.W., Tremblay, L.A., Chapman, H.F., 2005. Endocrine disruptors in

sewage effluent: the effects of formaldehyde preservation and the handling of sewage

samples. Proceedings of the Australian Water Association Contaminants of Concern in

Water Conference, June, Canberra, CD-ROM.

Leusch, F.D.L., Chapman, H.F., van den Heuvel, M.R., Tan, B.L.L., Gooneratne, S.R.,

Tremblay, L.A., 2006. Bioassay-derived androgenic and estrogenic activity in

municipal sewage in Australia and New Zealand. Ecotoxicol. Environ. Saf. 65, 403

411.

xxii

Chapter 1: Thesis objectives

1.1 General introduction

The endocrine system is diverse and complex, with varied and sophisticated

mechanisms that control hormone synthesis, release and activation, transport as well as

metabolism and delivery to the surface or interior of cells upon which they act

(Greenspan and Strewler, 1997). Endocrine disrupting compounds (EDCs) are defined

as exogenous substances or mixtures that can alter the function(s) of the endocrine

system and may cause health effects in an intact organism, or its progeny (WHO, 2002).

Industrial, agricultural and municipal wastes usually contain EDCs resulting in exposure

of organisms in the environment to unusually high concentrations of natural and

anthropogenic compounds that can elicit biological effects (Purdom et al., 1994;

Routledge and Sumpter, 1996). In wastewater, these compounds are sometimes able to

pass through the wastewater treatment system and reach receiving environments. It has

been demonstrated in Europe (Lavado et al., 2004; Diniz et al., 2005) and the USA

(Folmar et al., 1996; McArdle et al., 2000) that male fish held in treated wastewater

effluents or in rivers below wastewater treatment plants (WWTPs) showed a

pronounced increase of estrogen-dependent plasma vitellogenin concentrations. In egg-

laying vertebrates such as fish, estrogens activate the hepatic synthesis of vitellogenin.

This response has been suggested as a biomarker of exposure to estrogen active

substances (Sumpter and Jobling, 1995; Folmar et al., 1996).

As they are part of complex effluents, EDCs exist as mixtures. Individual compounds

within mixtures may vary greatly in estrogenic potency and may interact with each

other in an unpredictable manner. Measuring the concentrations of EDCs present in

water or solid phases typically involves extraction and analysis steps. It is essential to

develop effective methods that can extract multiple EDCs simultaneously from water

samples. Solid-phase extraction (SPE) is commonly used for spot or grab sample

extraction because of the large choice of sorbents for trapping targeted analytes. Aquatic

EDC monitoring programs are generally based on collection of discrete samples of

water phases. In environments where the contaminant concentrations may vary over

time, it is often desirable to expand the time window and increase the resolution by

taking more samples. Such pseudo time-integrated sampling of water, be it automatic or

manual, is both costly and cumbersome, and rarely used in large scale monitoring

studies. Passive sampling methods may represent a versatile tool in aquatic monitoring

1

programs, allowing a time-integrated monitoring of organic pollutants directly in the

aqueous phase as an alternative to conventional sampling techniques (Stuer-Lauridsen,

2005). During the past few years, miniaturization has become a dominant trend in

analytical chemistry with the development of stir-bar sorptive extraction (SBSE),

commercialized under the name Twister (Gerstel, Mlheim an der Ruhr, Germany). The

main advantages of this method are high sensitivity and a wide application range that

include extraction of volatile aromatics, halogenated solvents, polycyclic aromatic

hydrocarbons, polychlorinated byphenyls, pesticides, preservatives, odour compounds,

organotin compounds and EDCs from a variety of matrices (Tienpont et al, 2003;

Kawaguchi et al., 2005; Nakamura et al., 2005; Zuin et al., 2005; Duran Guerrero et al.,

2006).

Established analytical protocols are available for many of the compounds implicated as

being EDCs. Biological methods can also be used as screens to determine if EDC-active

compounds are present in a given environmental sample. In vitro and in vivo bioassays

offer a rapid, sensitive and relatively inexpensive solution to some of the limitations of

instrumental analysis. These bioassays can be used as tools to measure relevant

endpoints used for risk assessment of EDCs on the receiving environments. The

bioassays can be carried out concurrently with chemical methods to establish cause and

effect relathionships and to quantify the EDC activity present (Cech et al., 1998). New

and revised toxicological testing methods are being developed around the world

incorporating molecular and cellular biology and they hold promise for reducing whole

animal testing.

Several researchers have proposed and reported mathematical models which can be

used to quantify the distribution and fate of polycyclic aromatic hydrocarbons,

pharmaceuticals, pesticides, natural hormones and xenoestrogens in WWTPs (Clark et

al., 1995; Byrns, 2001; Khan and Ongerth, 2002 and 2004; Johnson and Williams,

2004). Clark et al. (1995) have modelled and analyzed the fate of organic chemicals in a

WWTP using fugacity modelling equations that describe the partitioning,

biodegradation, and volatilization or stripping behavior of chemical, which can be

solved to give an overall mass balance.

In South East Queensland, Australia, water levels in the major dams that supply potable

water to Brisbane are currently at an all time low of less than 30% of their maximum

2

capacities because of the sparse rainfall in the catchments areas over the past few years

(Figures 2.1 and 2.2). The state government has also proposed to add recycled water

from WWTPs into the regions dam as a measure to ensure the water level in the dam

does not fall below 10%. With this new proposal, there are concerns from the local

community over the presence of toxic substances, including EDCs, which might not be

fully removed by the WWTPs.

Currently, there are very few studies that address the removal efficacies of EDCs in

different treatment technologies of WWTPs in Australia. Furthermore, the comparison

between a variety of sampling techniques and analytical (chemical and biological)

results are rarely carried out in the various treatment trains of a WWTP to give an

overall EDC assessment of the plant removal efficacy and the effects of effluent

discharge have on the receiving environment. Within this context, this PhD research

presents several different analytical approaches to address the assessment of EDCs in

Australia.

1.2 Aims and objectives

The main aim of this research was to determine EDC concentrations and total estrogenic

activity in WWTPs in South East Queensland, Australia and to evaluate the practicality

of various collection and extraction methods. This was addressed using chemical

techniques to quantify the EDCs. In addition, a biological assay was utilized to predict

likely impacts in receiving environments. The five objectives of this research are listed

below:

(a) Development of suitable extraction technique for chemicals with endocrine

disrupting activity

The first objective was to develop a robust extraction technique that could extract

estrogenic compounds in wastewater and sludge with high recovery. Three methods

were used for this purpose; solid phase extraction (SPE) for grab sampling,

EmporeTM disk as the matrix for passive sampling and stir bar sorptive extraction as

a new extraction method for water and sludge (Chapter 3).

3

(b) Assessment of estrogenic compounds present in wastewater samples using

chemical analysis

Two new gas chromatography-mass spectrometry methods were developed to measure a

range of selected EDCs present in wastewater and sludge samples. Fifteen EDCs

were selected based on their potency and ubiquity in WWTPs and the receiving

environments. These compounds include the natural female and male hormones,

phthalates, alkylphenols and tamoxifen (Chapter 3, 4 and 5).

(c) Assessment of biological response using in vitro assay

The MCF-7 cell proliferation assay or E-Screen was used to determine the level of

estrogenic activity of the various wastewater samples (Chapter 6).

(d) Integration of chemical and biological techniques

Both chemical and biological assays were used to determine the estrogenicity of the

wastewater (influent and effluent) collected from 5 WWTPs in South East

Queensland, Australia. The comparison of efficacy of these 5 WWTPs at removing

estrogenic compounds and activity was assessed. Furthermore based on the results,

an estimation of hazard towards aquatic organisms was made based on the effluent

released into receiving environments (Chapter 5 and 6).

(e) Fugacity fate modeling for EDCs in a WWTP

Based on the selected EDCs concentrations in the WWTP, their fate was modelled

using a fugacity format with equations describing the partitioning, biodegradation,

and volatilization or stripping behavior of chemical, which can be solved to give an

overall mass balance. With this particular model, the various EDC removal

pathways from the WWTP can be identified (Chapter 7).

1.3 Research questions

i) Passive samplers allow time integrated evaluation of EDCs in WWTPs

Since passive samplers are time integrated and cost effective, it was predicted that

this technique would more easily provide ambient field EDC concentrations in

WWTP samples compared to the grab samples (Chapter 3, 5 and 6).

ii) Combining chemical and biological analyses will give a thorough interpretation

of estrogenicity

4

Chemical and biological analyses have been shown to have their own advantages

and disadvantages. It was predicted that combining the results from the chemical

and biological assays will provide a more complete understanding of estrogenic

activity and the compounds most likely responsible in order to trace the source of

the release or problem (Chapter 5 and 6).

iii) Estrogenic activity in WWTPs and receiving environments are caused by natural

hormones

Since natural hormones are in general more potent than industrial estrogen mimics,

it was predicted that even if there are trace amounts of estrogens found in

wastewater as compared to the high concentrations of industrial estrogen mimics,

the majority of estrogenic activity would be attributed to the natural estrogens

(Chapter 5 and 6).

iv) WWTPs significantly remove EDCs by the end of the treatment process

The activated sludge treatment process of a WWTP was predicted to be the most

effective step in biodegrading or removing a large portion of EDCs from wastewater

(Chapter 4, 5, 6 and 7).

v) EDCs fugacity fate modeling will provide a good understanding of EDC

removal

Using specific fugacity based equations, chemical physical properties and field

monitoring data, it was predicted that fate modeling of EDCs removal pathways in a

WWTP can be undertaken to reflect the ambient removal mechanisms (Chapter 7).

1.4 Thesis format

Except for Chapter 2, the chapters in this thesis are structured as stand-alone scientific

papers. This has led to some overlap in the material and methods section (particularly

between Chapter 3, 4, 5 and 6). Some material has been deliberately excluded from the

general introduction and literature review to avoid repetition in the introductions to data

chapters. The specific discussions in each chapter include most of the discussion

material, while a more concise general discussion at the end is aimed to highlight

synergies between the different chapters and to show the coherence of the overall

purpose of the research.

5

1.5 References

Byrns, G., 2001. The fate of xenobiotic organic compounds in wastewater treatment

plants, Water Res. 35, 2523 2533.

Clark, B., Henry, J.G., Mackay, D., 1995. Fugacity analysis and model of organic

chemical fate in a sewage treatment plant. Environ. Sci. Technol. 29, 1488 1494.

Diniz, M.S., Peres, I., Pihan, J.C., 2005. Comparative study of the estrogenic responses

of mirror carp (Cyprinus carpio) exposed to treated municipal sewage effluent

(Lisbon) during two periods in different seasons. Sci. Total Environ. 349, 129

139.

Duran Guerrero, E., Natera Marin, R., Castro Mejias, R., Garcia Barroso, C., 2006.

Optimisation of stir bar sorptive extraction applied to the determination of volatile

compounds in vinegars. J. Chromatogr. A 1104, 47 53.

Folmar, L.C., Denslow, N.D., Rao, V., Chow, M., Crain, A., Enblom, J., Marcino, J.,

Guillette, L.J., 1996. Vitellogenin inductions and reduced serum testosterone

concentrations in feral male carp (Cyprinus carpio) captured near a major

metropolitan sewage treatment plant. Environ. Health Perspect. 104, 1096 1101.

Greenspan, F.S., Strewler, G.J. (Eds.), 1997. Basic and Clinical Endocrinology. 5th

edition. Appleton and Lange, Stamford, CT, pp. 1 36.

Johnson, A.C., Williams R.J., 2004. A model to estimate influent and effluent

concentrations of estradiol, estrone and ethinylestradiol at sewage treatment works.

Environ. Sci. Technol. 38, 3649 3658.

Kawaguchi, M., Sakui, N., Okanouchi, N., Ito, R., Saito, K., Nakazawa, H., 2005. Stir

bar sorptive extraction and trace analysis of alkylphenols in water samples by

thermal desorption with in tube silylation and gas chromatography-mass

spectrometry. J. Chromatogr. A 1062, 23 29.

Khan, S.J., Ongreth, J.E., 2002. Estimation of pharmaceutical residues in primary and

secondary sewage sludge based on quantities of use and fugacity modelling. Water

Sci Technol. 46, 105 113.

Khan, S.J., Ongreth, J.E., 2004. Modelling of pharmaceutical residues in Australian

sewage by quantities of use and fugacity calculation. Chemosphere 54, 355 367.

Lavado, R., Thibaut, R., Ralda, D., Martn, R., Porte, C., 2004. First evidence of

endocrine disruption in feral carp from the Ebro River. Toxicol. Appl. Pharmacol.

196, 247 257.

6

McArdle, M., Elskus, A., McElroy, A., Larsen, B., Benson, W., Schlenk, D., 2000.

Estrogenic and CYP1A response of mummichogs and sunshine bass to sewage

effluent. Mar. Environ. Res. 50, 175 179.

Nakamura, S., Daishima, S., 2005. Simultaneous determination of 64 pesticides in river

water by stir bar sorptive extraction and thermal desorption-gas chromatography-

mass spectrometry. Anal. Bioanal. Chem. 382, 99 107.

Purdom, C.E., Hardiman, P.A., Bye, V.J., Eno, N.C., Tyler, C.R., Sumpter, J.P., 1994.

Estrogenic effects of effluents from sewage treatment works. Chem. Ecol. 8, 275

285.

Routledge, E.J., Sumpter, J.P., 1996. Estrogenic activity of surfactants and some of their

degradation products assessed using a recombinant yeast screen. Environ. Toxicol.

Chem. 15, 241 248.

Stuer-Lauridsen, F., 2005. Review of passive accumulation devices for monitoring

organic micropollutants in the aquatic environment. Environ. Pollut. 136, 503 524.

Sumpter, J.P., Jobling, S., 1995. Vitellogenin as a biomarker for estrogenic

contamination of the environment. Environ. Health Perspect. 103 (Suppl. 7), 173

178.

Tienpont, B., David, F., Benijts, T., Sandra, P., 2003. Stir bar sorptive extraction-

thermal desorption-capillary GC-MS for profiling and target component analysis of

pharmaceutical drugs in urine. J. Pharm. Biomed. Anal. 32, 569 579.

WHO, 2002. Global assessment of the state-of-science of endocrine disruptors.

Damstra, T., Barlow, S., Bergman, A., Kavlock, R., Van der Kraak, G. (Eds.),

International Program on Chemical Safety, World Health Organization.

Zuin, V.G., Montero, L., Bauer, C., Popp, P., 2005. Stir bar sorptive extraction and

high-performance liquid chromatography-fluorescence detection for the

determination of polycyclic aromatic hydrocarbons in Mate teas. J. Chromatogr. A

1091, 2 10.

7

Chapter 2: Literature review

2.1 Introduction The environment and organisms that live in it can be exposed to chemicals, including

those which may have endocrine disrupting activity, from such sources as agricultural

chemical use, industrial and commercial discharges to waterways and sewers as well as

excretion of natural and synthetic hormones by animals and humans to sewers. These

waters discharge, either directly or after treatment, to rivers or oceans. Human exposure

to chemical contaminants can be via food (naturally occurring contaminants, pesticide

residues, contaminants from transport or storage containers), through use of domestic

and consumer products (food packaging materials, pharmaceuticals products) and

potentially from drinking water.

Australia is a highly urbanised country, with its main population centres located on the

coastal fringe; and a limited number of smaller cities located inland. Thus, the bulk of

sewage effluent from the human population in Australia is treated and discharged to the

ocean. Agricultural runoff (including pesticides and fertilizers) from farming land in

the relatively small crescent of arable country running down the east coast into South

Australia, has the potential to find its way into creeks, streams and rivers which feed

into the Murray-Darling River system (Australias largest river catchment), the

Murrumbidgee River, or into a number of other rivers running east to the coast from the

Great Dividing Range. The Murrumbidgee and the Darling Rivers ultimately join the

Murray before it flows west, where it is used for irrigation and for drinking water.

Thus, in Australia, with respect to human health and exposure to endocrine disrupting

chemical contaminants in water, agricultural chemical runoff to rivers is likely to be of

greater concern than hormone discharge to city sewers (Falconer et al., 2003).

Because Australia is a dry continent, it has had to rely on very large reservoirs for the

supply of drinking water. In most States and Territories of Australia, these reservoirs

have highly protected catchments (e.g. Melbourne, Canberra, Sydney) and the water

supplies to their capital cities are of high quality. However, Adelaide, the capital of

South Australia, has to rely heavily on water taken from the Murray River, with the rest

of its supply obtained from reservoirs which have some agricultural land in their

catchments. Perth, the capital of Western Australia, relies on both reservoirs (with

protected catchments) and groundwater, for which there is the potential for

8

contamination from chemicals leaching into the sandy soil on which Perth is built.

Outside of the capital cities most country towns, apart from those located on large

rivers, rely on reservoirs which often collect from rivers and streams draining

agricultural catchments. Private dams in farming areas are quite likely to be

contaminated by agricultural runoff. Both dams and rainwater tanks in rural areas may

be contaminated if, for example, there is aerial spraying of crops.

In South East Queensland, Australia, water levels in the major dams that supply potable

water to Brisbane are

normal inactivation processes such as metabolism and excretion. That is, endocrine

disruption is not considered to be an adverse end-point per se, but rather is a mode or

mechanism of action potentially leading to other toxicological or ecotoxicological

outcomes e.g. reproductive, developmental, carcinogenic or ecological effects; these

effects are routinely considered in reaching regulatory decisions (at least for pesticides,

food additive chemicals and high production volume industrial chemicals for which the

required toxicology database is extensive).

In addition to endocrine disruption, there are other physiological mechanisms which can

be affected by excessive chemical exposure and chemical assessment should not unduly

focus entirely on carcinogens or endocrine disrupters but take into account all toxic end-

points of concern. Nevertheless, the focus on endocrine systems has led to an

acceleration of research and testing on a range of suspected problem chemicals and, in

many countries, has helped attract greater government and private funding for research.

10

Figure 2.1. Water supply catchments for nearly half of South East Queensland,

Australia region (SEQ Water, 2002).

11

s1065697Text BoxFigure removed, please consult print copy of the thesis held in Griffith University Library

Wivenhoe Dam

Figure 2.2. Water storage status of Wivenhoe, Sometset and North Pine Dams suplying

potable water to South East Queensland, Australia (SEQ Water, 2006).

2.2 The endocrine system The endocrine system and the nervous system are the major means by which the body

transmits information between different cells and tissues. This information results in the

regulation of most bodily functions. The endocrine system uses hormones to convey its

information. The endocrine system is diverse and complex, with varied and

sophisticated mechanisms that control hormone synthesis, release and activation,

transport as well as metabolism and delivery to the surface or interior of cells upon

which they act. Other mechanisms regulate the sensitivity of cells in target tissues to

hormones and the specific responses elicited by hormones. A hormone is defined as a

substance released by an endocrine gland and transported through the bloodstream to

another tissue where it acts to regulate functions of the target tissue (Greenspan and

Strewler, 1997). These actions are typically mediated by binding of the hormone to

receptor molecules. The receptor must be able to distinguish the hormone from a large

number of other molecules to which they are exposed to and transmit the binding

information to post-receptor events. Hormones are allosteric effectors that alter the

conformations of the receptor proteins to which they bind (Greenspan and Strewler,

1997).

Hormones produce their biological effects through interaction with high-affinity

receptors which are, in turn, linked to one or more effector systems within the cell. The

% F

ull

Somerset Dam North Pine Dam Average total system

12

effectors involve many different components of the cells metabolic machinery, ranging

from ion transport at the cell surface to stimulation of the nuclear transcriptional

apparatus. Steroids and thyroid hormones exert their effects in the cell nucleus, although

regulatory activity in the extranuclear compartment has also been documented. Peptide

hormones and neurotransmitters, on the other hand, trigger a plethora of signalling

activities in the cytoplasmic and membrane compartments while at the same time

exerting parallel effects on the transcriptional apparatus (Greenspan and Strewler,

1997).

2.3 Endocrine disrupting compounds (EDCs) It is now well established that there is a vast array of chemicals discharged into the

environment that can mimic (agonise) or block (antagonise) the action of hormones. A

hormone agonist is a compound that binds to a receptor and transmits binding into a

hormone response, while an antagonist is a compound that binds to a given receptor and

does not transmit the binding into a receptor response. The binding of an antagonist also

blocks binding of agonists and thereby prevents their actions, thus defining the term

antagonist. An endocrine disrupting compound is defined as an exogenous agent that

interferes with the synthesis, storage or release, transport, metabolism, binding, action

or elimination of natural blood-borne hormones responsible for the regulation of

homeostasis and the regulation of development process (Kavlock et al., 1996). Amongst

the important endocrine disruptors are those compounds suspected of interfering with

the normal action of the steroidal hormone estrogen through its receptor (i.e. estrogen

agonists and antagonists). The fact that these hormones play a critical role in the normal

development of the reproductive tract and sexual differentiation of the brain is well

documented (Cooper and Kavlock, 1997).

Disruption of sexual differentiation following exposure to estrogen has also been

demonstrated in various aquatic species, such as the turtle, which show temperature-

dependent sexual differentiation. Placement of either estrogen or some hydroxylated

polychlorinated biphenyls (PCBs) that are estrogen agonists directly on the egg have

been shown to alter sexual differentiation (Crews et al., 1995). Similar findings have

been reported in birds (Fry and Toone, 1981). Environmentally released chemicals may

also have anti-androgenic properties. Anti-androgenic compounds bind to the androgen

receptor, but block its transcriptional activity. Compounds such as the vinclozolin

metabolite M2 and the dichlorodiphenyltrichloroethane (DDT) metabolite, p,p-DDE,

13

inhibit androgen binding to the androgen receptor (Kelce et al., 1994 and 1995) and

androgen-induced transcriptional activity (Wong et al., 1995). In vivo studies of

vinclozolin and p,p-DDE have shown that these compounds inhibit androgen action in

developing, pubertal and adult male rats (Gray et al., 1994; Kelce et al., 1995).

Recently, concern has been expressed over the possibility that some synthetic chemicals

present in surface waters and aquatic sediments may adversely affect reproduction in

fish due to the possibility of pseudohermaphroditism and smaller testes weight (Purdom

et al., 1994; Sumpter, 1995).

2.4 Chemical properties of selected endocrine disruptors The exogenous chemicals in Table 2.1 with their molecular structures shown in Figures

2.3 2.7, have attracted much attention because even at low concentration levels they

are suspected of interfering with reproductive and behavioral health in humans and

wildlife, through disturbance of their endocrine system. Furthermore, these chemicals

are ubiquitous in the environment and have some of the highest potential as EDCs

compared to other known endocrine disruptors. From the physicochemical properties of

these compounds, it can be seen that most of them are in the range of low to moderately

hydrophobic organic compounds of mainly low volatility. It is expected that the

sorption on soil or sediment could be a significant factor in reducing their aqueous

phase concentration.

While reproductive toxicology studies in animals are typically required for regulation of

pesticides, many other chemicals in use have not been routinely screened for endocrine

disruption activity before being introduced for commercial use. Consequently the

significance of current concentrations of exposure to environmental estrogens or other

hormonally active compounds is unclear. To effectively assess the exposure to such

chemicals for endocrine disruption activity, the need for a rapid and sensitive screening

technique becomes apparent.

Because of the importance of the estrogen receptor (ER) in determining the

estrogenicity potential of a chemical which mimics or blocks the activity of natural

estrogens by specifically binding to the ER, there have been a number of attempts to

model the relationship between the structures of chemicals and estrogen receptor

binding affinity. Extensive binding studies of 17-estradiol analogues have indicated a

comprehensive binding property (Anstead et al., 1997; Brzozowski et al., 1997). That is,

14

15

the ER can bind with a wide variety of non-steroidal compounds, which are structural

analogues of the alkyl substituted phenol moiety of the 17-estradiol. For this steroid it

has been proposed that hydrogen bonding between the phenolic hydroxyl group and the

binding site in the ER, and also the hydrophobic and steric properties are important for

the binding affinity (Anstead et al., 1997; Brzozowski et al., 1997). For other

compounds, binding affinity depends on the extent of structural similarity with the

natural substrate.

One thing that has become very clear is the enormous difference in potency of

chemicals possessing estrogenic activity and probably other types of estrogenic activity

(Table 2.1). The most potent are the natural estrogens, such as 17-estradiol and the

synthetic estrogen 17-ethynylestradiol. Most, and perhaps all xenoestrogens (synthetic

chemicals that mimic the effect of estrogens) are much less potent, usually by 3 or 4

orders of magnitude, but sometimes even more. Thus, to obtain the same degree of

estrogenic response, it is usually necessary for the organism to be exposed to a much

higher concentration of xenoestrogen than that of 17-estradiol and 17-

ethynylestradiol. Obviously potency needs to be considered along with environmental

concentrations. Essentially all of the evidence to date suggests that it is the potent

steroidal estrogens that are the primary causative agents leading to feminization of fish

(Desbrow et al., 1998). Despite the general agreement that steroidal estrogens cause

much of the feminization of fish that has been reported, there appear to be at least a few

specific locations where concentrations of alkylphenolic chemicals, in particular

nonylphenol, are high enough that they contribute to the feminization, or may even be

the major causative chemicals (Sol et al., 2000; Sheahan et al., 2002; Todorov et al.,

2002).

16

Tabl

e 2.

1. B

ioch

emic

al p

rope

rties

of s

elec

ted

endo

crin

e di

srup

tors

.

Com

poun

d Ty

pe o

f com

poun

d M

olec

ular

w

eigh

t M

eltin

g po

int (

C)

Boi

ling

poin

t (C

) So

lubi

lity

in w

ater

(g

/100

mL)

Lo

g K

ow a

EE

q (e

stro

gen

equi

vale

nt) b

17-

estra

diol

N

atur

al st

eroi

d es

troge

n 27

2 17

3 -

1.0

10-

3 4.

01

1.0

(Dre

wes

et a

l., 2

005)

c 17-

estra

diol

N

atur

al st

eroi

d es

troge

n 27

2 17

3 -

1.0

10-

3 4.

01

0.10

(Kui

per e

t al.,

199

7) d

Estro

ne

Nat

ural

ster

oid

estro

gen

270

255

- 3.

0 1

0-3

3.13

0.

01 (L

eusc

h et

al.

2006

a) c

Estri

ol

Nat

ural

ster

oid

estro

gen

288

282

- B

arel

y so

lubl

e 2.

45

0.30

(Gut

endo

rf a

nd W

este

ndor

f, 20

01) c

17-

ethy

nyle

stra

diol

Fe

mal

e co

ntra

cept

ive

296

142

14

6 -

4.8

10-

4 3.

67

1.25

(Gut

endo

rf a

nd W

este

ndor

f, 20

01) c

Te

stos

tero

ne

Nat

ural

ster

oid

andr

ogen

28

8 15

2

156

- 3.

9 1

0-5

3.32

1

10-

5 (L

eusc

h et

al.

2006

a) c

Etio

chol

anol

one

Nat

ural

ster

oid

andr

ogen

29

0 18

1

184

-

Bar

ely

solu

ble

3.69

5

10-7

(Leu

sch

et a

l. 20

06a)

d A

ndro

ster

one

Nat

ural

ster

oid

andr

ogen

29

0 18

1

184

-

Bar

ely

solu

ble

3.69

5

10-7

(Leu

sch

et a

l. 20

06a)

d Ta

mox

ifen

Bre

ast c

ance

r tre

atm

ent

drug

37

2 96

9

8 -

OH

H

H

HOH

OHCH3 CH3 H Figure 2.3. Chemical structure of natural estrogens. Figure 2.4. Chemical structure of androgens.

17-Estradiol

HHOH

17-Estradiol

OH OCH3

H

H

HOH

CH3

H OH

Estrone

HHOH

Estriol

Testosterone

O

H

CH3

H

H

CH3

HOH

Androsterone

O

H

CH3

H

H

CH3

HOH

Etiocholanolone

O

CH3 H

H

CH3 OH

H

17

ON

Tamoxifen 17-Ethynylestradiol OH

H

H

H

CH3OH

CH

Figure 2.5. Chemical structure of pharmaceutical drugs.

OHCH3

CH2C(CH3)3CH3

OH C9H19 Nonylphenol 4-tert-octylphenol

CH3

Figure 2.6. Chemical structure of alkylphenols. Figure 2.7. Chemical structure of phthalates.

CH3 OH OH

CH3

OH

4-cumylphenol

CH3

Bisphenol A

O

O

O

OO

O

O

Diethyl phthalate

O

Dibutyl phthalate

CH3O

O

O

O

O

Benzyl butyl phthalate

O

O

O

CH2

CH2CH3

Di-(2-ethylhexyl) phthalate

18

2.4.1 Estrogens

The estrogens (17-estradiol, estriol and estrone) are predominantly female hormones,

which are important for maintaining the health of the reproductive tissues, breasts, skin and

brain. 17-Ethynylestradiol on the other hand is a synthetic steroid used as a contraceptive.

All vertebrate animals, including humans, can excrete steroidal hormone from their bodies,

which end up in the environment through sewage discharge and animal waste disposal. The

hormones 17-estradiol and estrone are naturally excreted by women (2 12 and 3 20

g/person/day, respectively) and female animals, as well as by men (estrone 5

g/person/day) (Gower, 1975). Pregnant women have been measured to excrete 260

g/person/day of 17-estradiol, 600 g/person/day of estrone and 6000 g/person/day of

estriol (Fotsis et al., 1980). However, Berg and Kuss (1992) demonstrated from a survey of

220 pregnant women that women could vary quite markedly in their excretions between

one another, and depending on the stage of their pregnancy. Based on the survey and

previous measurements of human estrogen excretion, Johnson et al. (2000) estimated the

daily excretion of estrogen by males and various categories of females (Table 2.2). From

such data on daily human excretion of estrogens, dilution factors and previous field

measurements, ng/L concentrations of estrogens are expected to be present in aqueous

environmental samples from English rivers (Johnson et al., 2000). These steroids have been

detected in effluents of sewage treatment plants and surface water (Ternes et al., 1999a).

They may interfere subsequently with the normal functioning and development in wildlife

(Jobling et al., 1998). Vitellogenesis (plasma vitellogenin induction) and feminization in

male fish have been observed in British rivers and are attributed to the presence of

estrogenic compounds (Desbrow et al., 1998; Jobling et al., 1998). Concentrations as low as

1 ng/L of estradiol led to the induction of vitellogenin (egg protein normally found in

female fish) in male trout (Purdom et al., 1994; Hansen et al., 1998).

In humans and animals, estrogens undergo various transformations, mainly in the liver.

They are frequently oxidized, hydroxylated, deoxylated or methylated prior to the final

conjugation with glucuronic acid or sulphate. 17-estradiol is rapidly oxidized to estrone,

which can be further converted into estriol, the major excretion product. Many other polar

metabolites such as 16-hydroxy-estrone, 16-ketoestrone or 16-epiestriol are formed and can

also be present in urine and faeces. The contraceptive ingredient mestranol is converted

after administration into 17-ethynylestradiol by demethylation (Ternes et al., 1999a). 17-

19

ethynylestradiol is mainly eliminated as conjugates, whereas other metabolic

transformations occur, but are of minor relevance. Therefore, estrogens are excreted mainly

as inactive conjugates with sulphate and glucuronic acid. Although steroid conjugates do

not possess a direct biological activity, they can act as precursor hormone reservoirs able to

be reconverted to free steroids by bacteria in the environment (Baronti et al., 2000; Ternes

et al., 1999a). Due to the presence of microorganisms in raw sewage and sewage treatment

plants, these inactive conjugates of estrogenic steroids are cleaved, and active estrogenic

steroids may be released to the environment (Baronti et al., 2000; Ternes et al., 1999a).

In an aerobic batch experiments with activated sludge, 17-estradiol was oxidized to

estrone, which was eliminated from the activated sludge tank without any further

transformation observed (Ternes et al., 1999b). The contraceptive 17-ethynylestradiol was

largely persistent under selected aerobic conditions, whereas mestranol was rapidly

eliminated and small portions of 17-ethynylestradiol were formed by demethylation. In

another experiment (Layton et al., 2000), 70 80% of added 17-estradiol was mineralised

to CO2 within 24 hours by biosolids from WWTPs, whereas the mineralization of 17-

ethynylestradiol was 25 75 fold less. 17-ethynylestradiol was also reported to be

degraded completely within 6 days by nitrifying activated sludge resulting in the formation

of hydrophilic compounds (Vader et al., 2000).

Table 2.2. Daily excretion (g) of estrogenic steroids by humans a.

Category 17-estradiol Estrone Estriol 17-ethynylestradiol Males 1.6 3.9 1.5 - Menstruating females

3.5 8 4.8 -

Menopausal females

2.3 4 1 -

Pregnant women 259 600 6000 - Women on contraceptives

- - - 35

a Estrogen concentrations taken from Johnson et al. (2000).

20

2.4.2 Tamoxifen

Considerable attention has been paid to the mechanism of action of triphenylethylenic

antiestrogens after they were demonstrated to antagonize the development of breast

cancers, especially those expressing the estrogen receptor . Among these drugs, the partial

anti-estrogenic tamoxifen has become a reference compound in view of its high clinical

efficacy and lack of major side effects (Favoni and de Cupis, 1998; Green and Furr, 1999;

Prichard, 2000; Plouffe, 2000). Tamoxifen has a non-steroidal triphenylethylene structure

which competes with estrogen for binding sites in the breast (Figure 2.5). At present anti-

estrogenic properties make tamoxifen the endocrine treatment of choice for all stages of

breast cancer. In addition, tamoxifen has a variety of other mechanisms which may mediate

its effect such as the induction of transforming growth factor from stromal fibroblasts,

the reduction in circulating levels of insulin-like growth factor I, inhibition of angiogenesis

and induction of apoptosis (Neven and Vergote, 2001).

Experimental studies conducted with the MCF-7 breast cancer cell line have clearly shown

that short term exposure to tamoxifen, as well as to its active metabolite 4-

hydroxytamoxifen, leads to a significant increase of ER content or up regulation (Kiang et

al., 1989; Gyling and Leclercq, 1990; Leclercq et al., 1992). Actually, additional

investigations with other partial anti-estrogens reveal that ER up regulation could be a

characteristic feature of this particular class of pharmacological compounds (Jin et al.,

1995; Legros et al., 1997). This behavior contrasts with that observed with other ligands

(i.e. estrogens, pure antiestrogens), which down regulate the receptor (Dauvois et al., 1993;

Devin-Leclerc et al., 1998).

ER up regulation upon tamoxifen treatment is associated with its strong anchorage to the

nuclear matrix (Oesterreich et al., 2000), which results in a progressive loss of 17-estradiol

binding ability (El Khissiin et al., 2000). The partial anti-estrogenicity of tamoxifen

suggests that this tamoxifen-receptor complex which is unable to bind with 17-estradiol

would not mediate transcription under an estrogenic stimulus while it may still respond to

signals generated by peptide growth factors (cross-talk mechanisms) (Lee et al., 2000;

Sakamoto et al., 2002). On the other hand, such ER accumulation does not seem to be

directly responsible for the cytostatic or cytotoxic effects of tamoxifen, since it is observed

in MCF-7 sublines resistant to high doses of this drug (Leclercq et al., 1992; Jin et al.,

21

1995). Up till now, there are still no studies reporting the impact tamoxifen has on the

environment and wildlife.

2.4.3 Androgens

Androgens (testosterone, etiocholanolone and androsterone) are predominantly male

hormones that stimulate or control the development and maintenance of masculine

characteristics in vertebrates by binding to androgen receptors. Androgen concentrations in

humans are generally much higher than estrogen concentrations. For example, plasma

testosterone concentrations are 3000 10,000 ng/L in adult males and 200 750 ng/L in

adult females, while 17-estradiol plasma concentrations are usually 10 60 ng/L in adult

males and 30 400 ng/L in adult females although they can be as high as 350 2000 ng/L

during pregnancy (Tietz, 1987). Kirk et al. (2002) reported that most of the androgenic

activity in municipal sewage with a predominantly domestic input is most likely caused by

androgens excreted by humans. Leusch et al. (2006b) found raw and treated wastewater

from WWTPs located in South East Queensland, Australia and New Zealand to have on

average 50 100 fold higher androgenic activity than estrogenic activity. Androgenic

activity in raw wastewater in the United Kingdom which ranged from 113 4300 ng/L

androgenic equivalents was also found by Kirk et al. (2002). As was the case with

estrogenic activity, WWTPs with activated sludge treatment were more effective than

trickling filters at removing the androgenic activity, with 82 99% net removal in activated

sludge plants compared to 57% in the tricking filter plant (Leusch et al., 2006b). Similar to

estrogens, sorption to activated sludge appears to be the major mechanism involved in

removing androgens from the aqueous phase (Esperanza et al., 2004; Layton et al., 2000).

Little is known about the effects of exposure of fish to androgenic chemicals. The lowest

observable effect concentration for induction of the male-specific protein, spiggin, in

female stickle backs (Gasterosteus aculeatus) after 3 5 weeks of exposure to

dihydrotestosterone was 2000 3000 ng/L (Katsiadaki et al., 2002), suggesting that fish

may not be susceptible to androgenic chemicals below the g/L concentration. However,

some studies have shown masculinization of mosquitofish exposed to paper mill effluents

containing ng/L concentrations of the steroid androstenedione (Ellis et al., 2003; Jenkins et

al., 2001).

22

2.4.4 Alkylphenols

Alkylphenol ethoxylates (APE) are a class of surfactants which are manufactured by

reacting an alkylphenol (e.g. nonylphenol and octylphenol) with ethylene oxide. An APE

molecule consists of two parts: the alkylphenol and the ethoxylate moiety. This structure

makes APEs soluble in water and helps disperse dirt and grease from soiled surfaces into

water. Alkylphenols have been found in various aquatic environments as products of

biological degradation of alkylphenol ethoxylates, which are used for a variety of industrial

applications due to their potential efficiency and low cost. Alkylphenols themselves are

also used as antioxidants and a stabilizer of plastics by some industries. Since alkylphenols

are more toxic, persistent, and estrogenic to aquatic living organisms than the ethoxylate

surfactants, the presence of alkylphenols in the environment has recently become of some

concern.

Alkylphenols such as nonylphenol, octylphenol, cumylphenol and bisphenol A, have been

shown to elicit estrogenic hormonal activity by binding specifically to estrogen receptors

(Soto et al., 1992; White et al., 1994; Hu and Aizawa, 2003). While there are significant

differences in the receptor-binding affinity of the various phenolic compounds, their

biological activity and the significance of exposure to them, even to those chemicals with

weak estrogenicity, they are nonetheless important because of their environmental

prevalence (Thiele et al., 1997). The structural feature responsible for the estrogenic

activity of alkylphenolic chemicals was found from the results of recombinant yeast

screening (Routledge and Sumpter, 1996). The estrogenicity is very dependant on the size

and degree of branching of the alkyl group, and its position on the phenol ring (Routledge

and Sumpter, 1997). The maximum response is found with eight carbons and a tertiary

branched structure. Other authors have also reported similar results (Taira et al., 1999, Blair

et al., 2000, Nishihara et al., 2000). The estrogenicity is dependent on the carbon number of

the straight chain alkyl group when the carbon number is less than seven.

Alkylphenols and APEs enter the environment primarily via industrial and municipal

WWTP effluent (liquid and sludge), but also direct discharge such as pesticide application.

The distribution of alkylphenols and their ethoxylates have been documented in many

studies in North America and Europe. Nonylphenol and octylphenol have been detected in

ambient air, water, soil, sediment and biota (Ying et al., 2002a).

23

Nonylphenol is widely used as plastic additive and antioxidant. A derivative of

nonylphenol, nonylphenol ethoxylate, is commonly used as a non-ionic surfactant in

detergents, paints, emulsifying agents, pesticides, herbicides as well as a dispersing agent

for industrial applications such as production of paper, fibre, metal and agriculture

chemicals (White et al., 1994, Nimrod and Benson, 1996; Khim et al., 1999). The in vitro

estrogenicity activity of nonylphenol was reported to be 10-6 times less than 17-estradiol at

a minimum (Jobling and Sumpter, 1993) to 2 10-3 times less at a maximum (Flouriot et al.,

1995). The no observed adverse effect level (NOAEL) for nonylphenol is 50 mg/kg body

weight (de Jager et al., 1999a and b).

Octylphenol is used for the production of octylphenol ethoxylates, a class of non-ionic

surfactants with a wide range of application. Octylphenol has been shown to weakly bind to

the estrogen receptor and to have weak estrogen-like activity in some in vitro screening

assays, with potency of octylphenol relative to estradiol of approximately 10-3 to 10-7(White

et al., 1994). In vivo screening assays have been variable with uterothropic responses and

other short-term changes occurring only at high doses, if at all (Gray and Ostby, 1998;

Williams et al., 1996). Madsen et al. (2002) found that a concentration of 4-tert-octylphenol

at 50 mg/kg body weight caused significant induction of vitellogenin in flounder

(Platichthys flesus).

Bisphenol A is a compound widely used as the monomer for the production of

polycarbonate plastic such as in baby bottles, and is a major component of epoxy resin used

for lining of food cans and dental sealants (Staples et al., 1998). To date, there have been

many reports detecting bisphenol A in the environment (Gonzalez-Casado et al., 1998,

Staples et al., 1998), baby food bottles (Mountfort et al., 1997), plastic waste (Yamamoto

and Yasuhara, 1998), and living organisms including humans (Miyakoda et al., 1999; Tan

and Mustafa, 2003). The safety of bisphenol A has become a controversial issue because it

not only possesses estrogenic endocrine disrupting effects (Krishnan et al., 1993; Brotons et

al., 1995), but also may be carcinogenic (Ashby and Tennant, 1988; Suarez et al., 2000).

There have been many reports concerning the disorders of reproductive organs when rats

and mice were exposed to bisphenol A in the prepubertal period (Vom Saal et al., 1998;

Stoker et al., 1999; Takao et al., 1999; Long et al., 2000; Tan et al., 2003). Bisphenol A was

able to activate estrogen receptors at concentrations lower than 1 M (Paris et al., 2002),

24

however the NOAEL for bisphenol A was set at 50 mg/kg body weight (Tyl et al., 2002). 4-

cumylphenol, just like bisphenol A, is commonly used in the manufacture of plastic

polymers and has been found to be a weak estrogen mimic (Hashimoto et al., 2001).

2.4.5 Phthalates

Phthalate esters are plasticizers used largely in the production of polyvinyl chloride

products to make them flexible and workable and, to a lesser degree, in paints, lacquers,

and cosmetics (Skinner, 1992; Harris et al., 1997). The physical rather than chemical

incorporation of phthalates in the polymeric matrix ensures that they are widespread

contaminants. Release of phthalates into the ecosystem or in wastewater effluents occurs

during the production phase and via leaching and volatilization from plastic products during

their usage and/or after disposal (Staples et al., 1997). Phthalates have been detected in

water, and air (Fatoki and Vernon, 1990). They have also been found in foods, especially in

fatty foods, as they can migrate out of food packaging materials (Sharman et al., 1994;

Petersen, 1991). Some phthalates are suspected of disrupting the endocrine system,

especially by mimicking estrogens (Harris et al., 1997). This assertion was primarily based

upon work conducted in vitro, using receptor binding assays or reporter cell systems, but

estrogenic activity was not a consistent finding.

The competitive binding of a phthalate to the estrogen receptor was first reported for

hepatic receptors derived from rainbow trout. Di-(2-ethylhexyl) phthalate (DEHP) did not

affect 17-estradiol at a concentration of 2 M, but there was a decrease in 17-estradiol

binding at higher concentrations, with a maximum of 25% reduction at 1 mM that was

suggestive of DEHP binding to the receptor. Dibutyl phthalate (DBP) was without effect at

a concentration of 80 nM, but induced a contraceptive-related decrease in 17-estradiol

binding at higher concentrations, with an apparent IC50 (the concentration required to

inhibit 17-estradiol binding by 50%) of 1 mM (Moore, 2000). Benzyl butyl phthalate

(BBP) inhibited 17-estradiol binding at all concentrations (estimated between 80 nM and

50 M), with an IC50 of approximately 10 M and maximum inhibition of 60% (Moore,

2000). Diethyl phthalate (DEP) displayed weak binding to Xenopus laevis liver cytosol,

with an IC50 of 12 M, representing a relative binding affinity of approximately 0.003

compared to 17-estradiol (Lutz and Kloas, 1999).

25

In an in vivo study, BBP, DBP, DEHP were assessed for estrogenic activity following

administration as four daily doses (20, 200, or 2000 mg/kg/day) to ovariectomized rats.

None of the phthalates stimulated either absolute or relative uterine weight increases in

immature animals, or vaginal epithelium cornification in mature animals (Zacharewski et

al., 1998a). In contrast, known estrogenic chemicals (including 17-estradiol) stimulated

uterine weight increase, vaginal cornification, and lordosis (Zacharewski et al., 1998a).

Benzyl butyl phthalate (BBP) is a phthalate ester that is present in paper and paperboards

used as packaging materials for aqueous, fatty, and dry food (IARC, 1982). BBP has been

tested for its estrogenic properties in vivo and in vitro. Uterothrophy and vaginal cell

cornification tests carried out on ovariectomized female Sprague-Dawley rats have shown

no estrogenic effects of BBP (Zacharewki et al., 1998b; Gray et al., 1999). In contrast, BBP

exerted estrogenic activities in several in vitro tests: MCF-7 cell proliferation, estrogen

receptor binding in rat uterus, and yeast transfected with human ER (Jobling et al., 1995;

Harris et al., 1997; Zacharewski et al., 1998b; Andersen et al., 1999).

2.5 Endocrine disruption

2.5.1 Mechanisms of endocrine disruption

A biologically active chemical can disrupt the endocrine system of an organism in a wide

variety of ways. The following are some examples, focusing particularly on the sex

hormone disruptors:

i) Binding to and activating the estrogen receptors (therefore acting as an estrogen) by

mimicking the female hormone 17-estradiol.

One complexity of this mode of action is the fact that there are a variety of estrogen

receptors, present in a wide range of tissues. It has been found that if several chemicals that

can bind and activate the estrogen receptor are added together, their effects will usually be

additive, so that effects of small quantities of a range of estrogenic chemicals can add

together into a much larger effect (Soto et al., 1995). Chemicals such as benzyl butyl

phthalate and di-n-butyl phthalate have been shown to add their effects to any natural

estrogen present (Jobling et al., 1995).

26

ii) Binding with but not activating the estrogen receptor (therefore acting as an anti-

estrogen).

For example, dioxin and furans work as anti-estrogenic agents through binding with the

aryl hydrocarbon receptor and estrogen receptor; however the aryl hydrocarbon ligand-

receptor complex may block estrogen receptor action in estrogen-responsive cells by DNA

binding competition (Krishnan and Safe, 1993; Klinge et al., 1999).

iii) Binding with other receptors.

There are many other receptors involved in the hormonal system, for example androgen

receptors for male hormones. This binding can either activate the receptor, or inactivate it,

as seen in anti-androgenic like effect of the DDT metabolite p,p-DDE (Kelce, 1995).

iv) Modifying the metabolism of natural hormones.

Some chemicals such as the pesticides lindane and atrazine, can affect the metabolic

pathway of estradiol, producing more estrogenic metabolites such as 16-hydroxyestrone,

potentially leading to an increased risk of breast cancer (Bradlow et al., 1995). Other

chemicals can activate enzymes which speed up the metabolism of hormones. The testes

contain specific enzymes to metabolise estrogens, breaking them down rapidly to a form

which can no longer bind to their receptor (Toppari et al., 1996). However, if these

enzymes are affected by a xenoestrogen, this metabolism will be reduced, increasing the

exposure of the testes to estrogen. This could be particularly relevant during fetal

development, when there are high concentrations of estrogen (Toppari et al., 1996).

v) Modifying the number of hormone receptors in a cell.

Complex mechanisms control the number of hormone receptors present in cells. A

chemical may reduce or increase the number of receptors, and so affect the existing

response to natural or synthetic hormones. For example, DDT and related compounds act

in a number of ways to disrupt endocrine function by binding with the estrogen receptor

(via mimicry and antagonism), altering the pattern of synthesis or metabolism of hormones

and modifying hormone receptor levels (Welch et al., 1969; Soto et al., 1995; Lascombe et

al., 2000; Rajapakse et al., 2001).

27

vi) Modifying the production of natural hormones.

Chemicals can affect natural hormone production by interfering with other signalling

systems, such as other hormone systems like the thyroid system, or the immune and

nervous systems. Chemicals such as pentachlorophenol affect the thyroid system by

reducing levels of thyroid hormone possibly through a direct