Cefas contract report C5108

Comparison of the Androgenised Female Stickleback Screen (AFSS) with other bioassays for detecting antiandrogens

Authors Ioanna Katsiadaki Marion Sebire

Issue date: 01/04/2011

Comparison of the AFSS assay with other bioassays for detecting antiandrogens

Authors:

Ioanna Katsiadaki (Contract manager, Cefas Weymouth) Marion Sebire

Issue date: 01/04/2011

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The AFSS and other bioassays for antiandrogens Executive Summary

1. The current OECD test guidelines for screening chemicals with potential endocrine disrupting activity (TG 229 and to a certain extend TG 230) cannot clearly identify androgen antagonists due to the lack of an androgen specific endpoint in the three validated species, the fathead minnow, the medaka and the zebrafish. 2. In mammalian toxicology there is a large literature body suggesting that chemicals with high affinity for the androgen receptor may be partly responsible for the increasing incidence of human reproductive disorders. An excellent mechanistic mammalian assay (the Hershberger assay) has played instrumental role in the characterisation of several chemicals as antiandrogenic. 3. The reported antiandrogenic activity in the aquatic environment is substantial. The evidence of high levels of antiandrogens in various water bodies emphasises the need for a suitable screening assay using fish. 4. Several fish species have secondary sexual characteristics whose development is controlled by androgens and therefore are promising as biomarkers for quantifying antiandrogen exposure. Of these, changes in anal fin morphology (gonopodium) in the mosquitofish and the guppy development of tubercles in fathead minnow, growth of papillary processes in the medaka, and induction of spiggin in the stickleback are discussed in detail. 5. Changes in the gonopodium in mosquitofish and guppy have not been exploited as a biomarker for antiandrogens to date (only as a marker for androgen exposure), the tubercle score in fathead minnow has produced variable responses in the literature whilst the studies on the effect of antiandrogens using as an endpoint the papillary processes in the medaka are limited. 6. In contrast, spiggin, the androgen biomarker, present in the kidney of the three-spined stickleback has been extensively used in this field, displaying an unprecedented sensitivity, specificity and resolution (dynamic range) for the detection of androgens and antiandrogens in the environment. 7. The in vivo test that was developed to specifically detect (anti)androgenic activity of chemicals using spiggin, the Androgenised Female Stickleback Screen (the AFSS, currently a draft guidance document) has undergone both validation and peer review and has proved its reproducibility and sensitivity, including lack of interferences from endocrine negative chemical(s).

The AFSS and other bioassays for antiandrogens

8. Currently the only Test Guideline validated by the OECD as a screen for (anti)androgens is the castrated rat Hershberger bioassay. The principle of the AFSS is very similar to the Hershberger rat assay; in both assays the androgen sensitive tissues are stimulated by an androgen. Simultaneous administration of putative antiandrogenic chemicals can block this effect providing excellent mechanistic models for the evaluation of chemicals with affinity for the androgen receptor(s). 9. The main difference between the two assays is that in the male rat, the endogenous source of androgens (testes) is eliminated by surgical castration whilst in the stickleback intact female fish are used instead of male fish, bypassing the need for surgical intervention. 10. In the AFSS the only measured endpoint is the kidney levels of spiggin; this protein is measured by ELISA (just like the fish vitellogenin assays) providing an impressive dynamic range for detecting antiandrogens. The difference a stimulated and a non-stimulated female kidney is about 1000-fold whilst in the Hershberger the differences in the weight of the androgen-sensitive tissues between stimulated and non stimulated rats is only about 5-fold. 11. Comparing the sensitivities of a mammalian and a fish in vivo test is a challenge; due to the different modes of chemical exposure the bioavailability of chemicals is dramatically different; the chemicals in the Hershberger are administered by injections (androgen) or oral gavage (antiandrogens) whilst in the AFSS exposure to both chemicals is waterborne. 12. Qualitative comparison of the AFSS and Hershberger data on seven chemicals demonstrated not only that both assays detected successfully their antiandrogenic potential but also their ranking in terms of potency was very similar.

The AFSS and other bioassays for antiandrogens

Table of contents

1 Introduction ...... 1 1.1 Background information ...... 1 1.2 The issue of environmental antiandrogens ...... 2 1.3 Scope of this report ...... 5 2 Androgen-dependent traits in fish ...... 6 2.1 General androgen-regulated traits in fish ...... 6 2.1.1 Changes in phenotypic sex ratio ...... 6 2.1.2 Reproductive behaviour ...... 7 2.1.3 Fecundity ...... 8 2.1.4 Vitellogenin ...... 8 2.1.5 Gonadal histology ...... 9 2.1.6 Summary of generic effects as markers for antiandrogens ...... 10 2.2 Species-specific androgen-regulated traits in fish ...... 11 2.2.1 Spiggin in the stickleback ...... 11 2.2.2 Gonopodium in mosquitofish and guppy ...... 13 2.2.3 Tubercles in the fathead minnow ...... 14 2.2.4 Anal fin (papillary processes) in medaka ...... 17 3 The AFSS and other bioassays for antiandrogens...... 18 3.1 The principle of the AFSS ...... 18 3.2 Alternative androgenised female fish for detecting antiandrogens ...... 19 3.3 The principle of the Hershberger bioassay ...... 20 3.4 Antiandrogen testing in the AFFS and the Hershberger assays ...... 21 4 Overall discussion and conclusions ...... 25 5 Tables ...... 27 Table 1. Synopsis of in vivo fish exposures to well known antiandrogens...... 28 Table 2. Synopsis of in vivo fish exposures to well known antiandrogens where females are stimulated by exogenous androgens...... 38 Table 3. Comparison of the Hershberger and AFSS assays in term of responses to antiandrogenic chemicals...... 40 6 References ...... 42

The AFSS and other bioassays for antiandrogens

List of abbreviations

11KT 11-ketotestosterone EE2 17 α-ethinyl-oestradiol E2 17β-oestradiol AR androgen receptor BPA bisphenol A COWS Cowper’s glands CA cyproterone acetate DDE dichloro-diphenyl-dichlorethylene DHT dihydrotestosterone ELISA enzyme linked immunosorbent assay FHM fathead minnow FN fenitrothion FSDT Fish Sexual Development Test FL flutamide GLANS glans penis KEH kidney secondary proximal epithelium height LABC levetor ani-bulbocavernosus muscle LN linuron LOEC Lowest Observed Effect Concentration MED medaka MT methyltestosterone NOEC No Observed Effect Concentration PCR polymerase chain reaction PRO procymidone SSC secondary sexual characteristics SV seminal vesicles plus coagulating glands T testosterone TP testosterone propionate SB three-spined stickleback TB trenbolone VP ventral prostate VZ vinclozolin VTG vitellogenin YAS yeast androgen screen ZEB zebrafish

The AFSS and other bioassays for antiandrogens

The AFSS and other bioassays for antiandrogens

1 Introduction

1.1 Background information In 1996, the US Environmental Agency (EPA) set up an Endocrine Disruptor Screening and Testing Advisory committee (EDSTAC) to establish screening programs for assessing the endocrine disrupting effect on humans and wildlife of over 85,000 chemicals in use (EDSTAC 1998). Similarly, the OECD endocrine disruptor activity was initiated in November 1996 and a task force on endocrine disrupter testing and assessment (EDTA) was established in 1997 to oversee the programme of method development for EDC detection (Huet, 2000). The initial framework consisted of three tiers - initial assessment, screening and testing. The OECD Secretariat subsequently established a Validation Management Group for Ecotoxicological Test Methods for Endocrine Disrupters (VMG-Eco), whose aim was to supervise the work on validation of EDC fish tests as well as tests using all other environmental species of concern. In fish species that have been used in the validation work for screening and testing endocrine disrupters are the fathead minnow ( Pimephales promelas ), the zebrafish ( Danio rerio ), and the medaka ( Oryzias latipes ). The three-spined stickleback ( Gasterosteous aculeatus ) participated successfully only in phase 1b of the validation programme hence it is not considered a fully validated species in the test guidelines for the screening of endocrine disrupting chemicals in fish, namely TG229 and TG230. The main shortfall of TG230 and to a certain extent of TG229 is the lack of ability to detect antiandrogens (OECD report, 2006), an important class of EDCs as detailed in section 1.2. Hence, currently the only robust and validated by the OECD program in vivo assay for antiandrogens is the Hershberger bioassay (TG441, adopted September 2009). In view of the numerous clinical implications for human health, the high antiandrogenic activity detected in the aquatic environment and the fact that the only reliable antiandrogen bioassay requires castrated rats, an in vivo test using intact fish in order to screen and identify antiandrogenic chemicals was sought as highly desirable. In this respect, the UK proposed that the Androgenised Female Stickleback Screen (AFSS) can fill this gap and submitted a Standard Protocol Submission Form (SPSF) form in May 2008. The work that comprised of a) a retrospective validation of a large dataset produced over 7 years of research, b) investigations into the effect of solvent and c) testing of negative compounds was commissioned in early July 2008. A total number of 20 exposures using the AFSS had already been conducted under Defra-funded research and the EU-funded project EDEN in 4 laboratories. The test design is

The AFSS and other bioassays for antiandrogens 1

similar to the OECD TGs 229 and 230 (with some minor differences) and so the data highly comparable to the validation studies of these TGs. Independent statistical analysis of a large dataset using the AFSS has revealed that the assay is robust and can detect antiandrogens unambiguously (OECD, 2010). Although the absolute levels of kidney spiggin levels were variable between the exposures, every single test successfully detected the antiandrogenic activity of the tested compounds and each set of data followed a dose response curve. Two candidate negative control chemicals were also tested. Potassium permanganate that was proved very toxic to fish (and hence its use as a negative was not supported) and ammonia, which proved a suitable endocrine negative compound with no effect in spiggin (either by means of spiggin induction or reduction in the treated female fish). The validation report has undergone a successfully peer review process and the AFSS is currently a draft guidance document. This report provides a comparative review of in vivo exposures of fish to antiandrogenic chemicals. In addition, we attempt a qualitative comparison of the AFSS responses to all tested antiandrogens to date with the data produced during the validation studies of the castrated rat Hershberger assay. .

1.2 The issue of environmental antiandrogens During the past decade an increase in the incidence of reproductive disorders has been reported in several animal taxa including humans. Several field and laboratory studies have investigated the putative links between these disorders and the environmental presence of anthropogenic chemicals that are able to interfere with the endocrine functions in animals (reviews by Depledge and Billinghurst, 1999; Oehlmann and Schulte-Oehlmann, 2003; Matozzo et al , 2008, Gray et al , 1998; Fossi and Marsili, 2003; Brunström et al , 2003; Crain and Guillette, 1998; Guillette and Iguchi, 2003; Kime et al , 1999; Jobling and Tyler, 2003; Caldwell et al , 2008). In teleost fish the main focus of research for more than a decade has been on oestrogenic xenobiotics. The presence of the female-specific egg yolk protein precursor, vitellogenin (VTG) in male fish has been used as an unambiguous biomarker for oestrogen exposure in a variety of laboratory species whilst the widespread sexual disruption in wild fish was also associated with oestrogenic chemicals (Jobling et al , 2006). Besides, oestrogenic activity has been identified in a large number of domestic sewage and industrial effluents and the responsible chemicals have been identified through toxicity identification evaluation. The nature and potency of xenoestrogens identified to date display a large diversity and includes natural products, pesticides, pharmaceuticals, personal care products and industrial chemicals. However, feminisation of wild male fish is thought

The AFSS and other bioassays for antiandrogens 2

to be predominantly due to steroidal estrogens of human origin (natural and synthetic) present in sewage effluents with other estrogenic chemicals found in domestic effluents such as bisphenols, some phthalates, nonylphenols and their ethoxylates contributing only mildly to this phenomenon (Jobling et al , 2006). In human health, epidemiological and clinical studies attempted a link of environmental chemicals and reproductive disorders such as the increased incidence of idiopathic hypospadias, alterations of the male genitalia, pseudo-hermaphroditism, decline in sperm counts and the increasing incidence of breast cancer amongst females and testicular and prostate cancers amongst males (Skakkebaek and Keiding, 1994; Toppari et al , 1996; Irvine et al , 1996; Toppari and Skakkebaek, 1998; Paulozzi, 1999; Sharpe and Skakkebaek, 2003; Darbre, 2006; McLachlan et al , 2006). The potential risks associated with exposure to endocrine disrupting chemicals (EDCs) and human reproductive health however remains a controversial issue (Safe, 2000; Sultan et al , 2001; Sharpe and Skakkebaek, 2003; Sharpe and Irvine, 2004; Fisher, 2004; Tsutsumi, 2005; Vidaeff and Sever, 2005; Waring and Harris, 2005; Foresta et al , 2008). Nevertheless, it has been experimentally established that, in addition to oestrogen exposure that can effectively ‘feminize’ males, exposure to chemicals that antagonise the effect of androgens, directly inhibit the development of male characteristics (Kelce and Wilson 1997), cause de- masculinisation and reduce sperm production (Lambright et al , 2000; Gray et al , 2001; Williams et al , 2001; Daxenberger 2002; Hotchkiss et al , 2004; Gray et al , 2006). Androgens regulate a variety of tissue and species-specific genes. In mammals, testosterone (T) is inducing differentiation of the Wolffian ducts whilst its metabolite 5α-dihydrotestosterone (DHT) is responsible for the formation of external genitalia and testicular descent. Androgens are also critical during adulthood, maintaining reproductive function, behavior and secondary sexual characters. Both mammalian androgens, T and DHT, mediate their biological effects predominantly through binding to the Androgen Receptor (AR), an androgen-inducible member of the nuclear receptor superfamily of transcription factors. In teleost fish as in mammals, androgens are essential as they control male sexual differentiation, maturation and behaviour (Borg, 1994). They also act by binding to specific androgen receptors (ARα and ARβ) to activate or repress the expression of specific genes, notably involved in the development of male primary and secondary sexual characters (reviewed by Nagahama and Delvin, 2002). Although it is generally accepted that the physiological relevant androgen in fish is 11-ketotestosterone (11KT), the discovery of a nuclear receptor with high binding affinity for 11KT has not been achieved to date.

The AFSS and other bioassays for antiandrogens 3

In most teleost fish however, the mediation of androgen action is more complicated than in mammalian vertebrates as there is a duality in the active androgens involved in reproduction; both mammalian type of androgens (T and DHT) and 11-oxygenated androgens (i.e. 11KT) are functional androgens in fish (Borg, 1994; Margiotta-Casaluci and Sumpter, 2001). The amino acid sequences of both the DNA binding domain and the ligand binding domain are highly conserved from fish to mammals with 90% and 70% identity with mammalian ARs for the DBD and the LBD respectively (Touhata et al , 1999). Several duplicates of AR have been characterized in many Actinopterygian fish to date (review by Douard et al , 2008). These two receptors display different tissue distributions and different binding specificities suggesting that the receptors mediate the actions of different androgens in different tissues of teleost fish. Besides the fact that the very first example of chemical endocrine disruption that led to population crashes of gastropod molluscs in the 80’s was due to masculinisation (Bryan et al, 1986; Matthiessen et al , 1995), the potential impacts of environmental androgens and antiandrogens in fish sexual disruption has been generally overlooked for many years. Historic and new field data suggest that the presence of androgenic xenobiotics is restricted to pulp mill effluents (Howell and Denton, 1989; Larsson et al , 2002), feedlot (Wilson et al , 2002) and a low number of sewage effluents (Thomas et al , 2002). However, more recent reports suggest that compounds with antiandrogenic activity appear to be more widespread: In 2003 and 2004, two nationwide surveys revealed a significant antiandrogenic activity present in UK final sewage effluents (Johnson et al , 2007). More recently, significant antiandrogenic activity was reported in water and sediment samples from a river in Italy (Urbatzka et al , 2007) and in the produced water from oil platforms in the North Sea (Tollefsen et al , 2007). All three studies used the same, well-established in vitro method, namely the yeast androgen screen (YAS), in which the human androgen receptor is incorporated into the yeast cell. The recorded antiandrogenic activity in the environment is expressed in flutamide (FL) equivalents (FL eq), FL being the most commonly used model antiandrogen. Although the number of reports is so far small, the actual antiandrogenic activity is very significant; up to 1230µg/L FL eq in the UK final sewage effluent surveys, up to 4200µg/L FL eq in the river Lambro and an extraordinary 8000µg/L FL eq in the oil platform produced water. To date the identification of chemicals present in domestic sewage effluents, causing this high antiandrogenic activity in the YAS assay has been limited to two chemicals; dichlorophene and dichloromethylanthracene can account for a small fraction of the antiandrogenic activity of these effluents (between 2 and 19%) indicating that the vast majority of androgen antagonists in the aquatic environment is still unknown (Hill et al , 2010). In the marine environment the positive

The AFSS and other bioassays for antiandrogens 4

identification of naphthenic acids as the source of the extraordinary high antiandrogenic activity in the YAS has been recently reported (Thomas et al , 2009).

1.3 Scope of this report The newly emerged problem of high antiandrogenic activity in the aquatic environment and a recent desk study where statistical analysis of chemical and biological data suggested that antiandrogens are contributing to wild fish sexual disruption in UK rivers (Jobling et al , 2009) has highlighted the importance of a fish screen to detect this class of EDCs. The association between fish feminisation and presence of chemicals with androgen receptor antagonistic activity in the environment (albeit not experimentally proved) challenged the so far assumption that fish feminisation is mainly due to steroidal oestrogens. Chemical-induced disruption of the androgen signalling pathway in mammals can take place in different parts of the pathway eliciting the same phenotype via reduction of the androgen-depended gene expression and protein synthesis. For example, it is widely accepted that phthalates (a group of chemicals that received a lot of attention in mammalian toxicology) inhibit foetal testosterone synthesis and do not act via the AR whilst pesticides like vinclozolin for example bind to the AR, inhibiting transcriptional activity induced by natural androgens. This document will focus on discussing the potential of the AFSS and other relevant fish bioassays in detecting androgen receptor antagonists rather than other modes of action that chemical may have eliciting an overall antiandrogenic effect. This is not only because the AFSS is a highly specific mechanistic assay for detecting AR antagonists but also because the antiandrogenic activity in the aquatic environment appears to be due to chemicals that bind to the AR. Numerous in vivo mammalian studies have reproduced a demasculinised phenotype by exposing laboratory rodents to antiandrogens and have indicated that various chemical classes display such activity, including pesticides and herbicides such as fenitrothion (FN), vinclozolin (VZ), procymidone (PRO), linuron (LN), DDT and its metabolite, p,p' ,-DDE. Most of these antiandrogenic chemicals have been detected by the Hershberger rodent assay (or one of its variants) which is a highly specific mechanistic assay for the detection of chemicals that display affinity for the androgen receptor. A qualitative comparison of the AFSS and Hershberger data on chemicals that were commonly tested is also part of the scope of this document. In fish there is currently a lack of a regulatory test in which antiandrogens are detected with certainty (OECD 2006b; Bars et al , 2011) besides the fact that numerous fish species display a variety of androgen regulated traits. A short review of these traits follows along with our assessment of their utility as a screen assay for detecting environmental antiandrogens.

The AFSS and other bioassays for antiandrogens 5

2 Androgen-dependent traits in fish

2.1 General androgen-regulated traits in fish The physiological role of androgens in teleost fish development, sexual maturation and behaviour is multifaceted and includes regulation of a diverse range of traits that could potentially be used as markers of androgenic endocrine disruption. Some of these are universal (i.e. they are commonly detectable in more than one species) and some are species-specific. Equally, some of these traits are specific to androgens whilst others respond to a diverse range of hormones and chemicals. In this section we limit our detailed discussion to the available data from short term in vivo fish exposures to known antiandrogenic chemicals. In addition, particular emphasis is given in studies that utilised the fathead minnow (FHM) and the medaka (MED) due to their important role in regulatory testing. The zebrafish (ZEB) lacks secondary sexual characters that can be deployed as androgen biomarkers hence its usefulness for the detection of this class of EDCs is limited. The responses of these species reported in the literature are compared with the stickleback responses to androgenic xenobiotics. A range of endpoints or combination of endpoints has been suggested as diagnostic for screening potential (anti)androgenic chemicals, including effects on reproductive behaviour, gonadal histology, vitellogenin, plasma steroid levels, phenotypic sex ratios, fecundity, egg fertility and gene expression profiling. Although the majority of these endpoints are apical and hence useful for risk assessment of chemicals, they suffer from lack of specificity as they can be affected by a diverse range of chemicals, not specifically androgenic xenobiotics. Table 1 (section 5) displays a synopsis of the fish studies employing antiandrogenic chemicals. For brevity and relevance, we included studies that utilised the three core OECD species (FHM, MED, ZEB), the stickleback (SB) studies (excluding the AFSS protocol) and other species only if an androgen-specific endpoint was used or the exposure utilised a model (not environmental) antiandrogen such as FL and cyproterone acetate (CA).

2.1.1 Changes in phenotypic sex ratio Chemical-induced alteration of phenotypic sex ratio (skewed sex ratio) is an important endpoint in testing and assessment; the expectation being that androgenic chemicals will skew the ratio towards males and antiandrogenic chemicals towards females. Nevertheless, the latter has not been experimentally achieved in the vast majority of short-term exposures (Makynen et al , 2000; The AFSS and other bioassays for antiandrogens 6

Kiparissis et al , 2003; Wester et al , 2003; Kristensen et al , 2006). Besides, skewed sex ratio is a core endpoint in the Fish Sexual Development Test (FSDT), a draft test guideline where exposure takes place soon after fertilisation and continues until 60 days post hatching, in order to include the critical period of sexual differentiation. In this respect it is unlikely that a short screening assay can exploit further this important and population relevant endpoint.

2.1.2 Reproductive behaviour We believe that effects on reproductive behaviour are one of the most interesting endpoints for chemical risk assessment. However, besides the fact that the role of androgens in programming the brain and determining successful reproductive behaviour is critical and well-characterised in many vertebrate species, the effects of antiandrogens on fish behaviour are not well-studied. The few studies that employed antiandrogenic chemicals used changes in the stickleback (Rouse et al , 1977; Sebire et al , 2008; Sebire et al , 2009) and the guppy (Baatrup and Junge, 2001; Bayley et al , 2002; Bayley et al , 2003; Kristensen et al , 2006) reproductive behaviours as measures of endocrine disruption. In three studies, male adult guppies (Poecilia reticulate ) were exposed to three antiandrogenic compounds, VZ, p,p’ -DDE, and the model antiandrogen FL via the food (Baatrup and Junge, 2001; Bayley et al , 2002; Bayley et al , 2003). The male guppy courtship was disturbed, with the sigmoid display, a suitable measure of the male’s mating ardour, affected by all chemicals (although not always in a dose response manner). In another guppy study however, p,p’ -DDE failed to reduce the reproductive fitness of males including courtship behaviour (Kristensen, 2006). The behaviour of the male stickleback goes through three distinct well-characterised phases (nest building, courtship, and parental care) the first two of which are linked to high concentrations of androgens in the blood (Wootton 1976; Borg and Mayer 1995; Rowland et al , 1995; Páll et al , 2002a; 2002b); scientists have been studying the stickleback reproductive behavior for over a century forming an extensive literature. Exposure of intact male sticklebacks to FL reduced severely the number of males that built a nest at 100µg/L and completely inhibited nest building behaviour at both at 500 and 1000µg/L (Sebire et al , 2008). Exposure of male sticklebacks to CA for 21 days (Rouse et al , 1977) resulted in a significant delay of the start of the nest building phase. In addition, exposure of male sticklebacks to FN significantly reduced nest-building activity and courtship (Sebire et al , 2009). Impairment of the reproductive behaviour in fish is a promising avenue for the development of an (anti)androgen bioassay due to its simplicity, cost effectiveness and population effect relevance. There is however a major drawback associated with this apical endpoint; it is not specific to antiandrogenic xenobiotics. Oestrogenic chemicals for example caused the same delay in the onset of nest building in the stickleback (Wibe et al , 2002; Brian et al , 2006). The AFSS and other bioassays for antiandrogens 7

2.1.3 Fecundity Similarly, although many antiandrogens affect fecundity, this endpoint is by no means diagnostic or specific to a mode of action directly linked to the androgen signalling pathway. Fecundity decline is neither consistent with antiandrogen exposure in fish (see table 1, section 5) and in the literature is more often associated with reproductive toxicity rather than endocrine disrupting potential. Interestingly, the conclusion of the peer review results for the fish short-term reproduction assay using the fathead minnow states that fecundity together with secondary sex characteristics seem to be the most consistent endpoints responding to androgen receptor antagonists such as FL and VZ (Easter Research Group, 2008). Although this may have been the case for VZ (Biever 2007, see table 1, section 5) the data from the follow-up inter-laboratory study on FL do not support this statement since one of the two laboratories did not detect any changes in fecundity even at the highest FL concentration (1000µg/L) tested (Battelle 2006, table 1, section 5). The large variability displayed by individual fish in sex steroid titres is well-known amongst toxicologists (see also table 1, section 5) and for this reason is not included as a mandatory endpoint in TGs 229 and 230. It also presents further drawbacks such as the requirement of highly specialised equipment or reagents for their analysis. Critically, in small teleost fish like those used for regulatory testing the amount of plasma available is often a limiting factor for this analysis as it is required for determining VTG.

2.1.4 Vitellogenin Alterations in VTG titres in fish have been suggested as a suitable endpoint for both androgenic (reduction in female VTG titres, induction in males by aromatisable androgens) and antiandrogenic (increase in female VTG, occasional increase in male VTG) exposures. This endpoint however, has demonstrated a large variability in the literature with the majority of the antiandrogenic exposures demonstrating no effect in either male or female VTG (see table 1, section 5). For example, a significant VTG induction was reported in female fathead minnows exposed to FL in all three concentrations (100, 320 and 1000µg/L) tested (Panter et al , 2004), but this was not confirmed by the validation work on the fish screen (OECD 2006). Similarly, adult medaka exposed to FL at a range of 90.4 and 1470µg/L (measured concentrations) showed no response in terms of VTG (Nozaka et al , 2004) but a significant increase was found in female medaka exposed to FL via the diet at 0.2mg/g of food (Chikae et al , 2004). Jensen et al (2004) suggested that the FL-induced increase in female VTG could be due to either the increase in plasma oestradiol (E2) and/or decreased deposition of the yolk protein in the oocyte. In this study, FL reduced fecundity of the fish (by reducing frequency of spawning), indicating that FL caused a delay in egg maturation and/or release. Since female fish are not able to clear plasma VTG by depositing in the developing egg, the result is a build-up of VTG in The AFSS and other bioassays for antiandrogens 8

female plasma. Although this may be a valid explanation, the increase in female VTG after exposure to antiandrogens is an endpoint that have showed high variability and as such we proposed that changes in VTG is not an suitable biomarker to detect androgens and antiandrogens, without defining a series of other endpoints such as fecundity, gonadal histology and sex steroid titres.

2.1.5 Gonadal histology Alteration of gonadal histology has also been proposed as an appropriate endpoint for detecting androgenic xenobiotics. In the context of a short term fish screen however this endpoint has also demonstrated high variability (see table 1, section 5). In medaka, one FL exposure resulted in no change in histology in either male of female fish (Kang et al , 2006) whilst two further studies reported no effects in either sex (OECD, 2006). Effects on female gonadal histology (decrease in oocyte size and maturation) have been reported in two medaka studies (OECD 2006) but only one out of the four laboratories participating in the validation work of the fish screening assay reported effects on testicular histology (advanced testicular score). Similarly, gonadal histopathology in fathead minnow studies with androgen antagonists demonstrated high variability. During the validation of the fish screen using FL, one of the three participating laboratories reported no changes in either sex, one laboratory reported effect in females (increased proportion of atretic oocytes) whilst one laboratory detected effects on male gonadal histology by means of an increased proportion of spermatogonia and decreased proportion of spermatocytes (OECD, 2006b). Other studies have not consistently reproduced the same results; Jensen et al , (2004) reported degeneration and necrosis in testicular histology as a result of FL in the FHM. Nevertheless, increased proportion of atretic oocytes was a common effect of various antiandrogens such FL (Jensen et al , 2004; Batelle, 2003), VZ (Martinovic et al , 2008) and p,p’-DDE (Battelle, 2005). In contrast, Makynen et al , (2002) when exposed fathead minnow embryos to VZ at concentrations ranging from 90-1200μg/L did not detect any adverse effects on sexual differentiation or reproductive health and when exposed adult fathead minnows to VZ (200 or 700 µ/L for 21 days) reported a marked reduction in gonadal condition of female fish at the high concentration only. The possibility that VZ and its metabolites would bind to the fathead minnow AR was investigated through competitive radioligand binding studies (Makynen et al , 2000). In this study, VZ and its metabolites failed to compete for high-affinity, low-capacity testosterone binding sites in FHM brain and ovary cytosolic fractions, suggesting that these chemicals might not act as antiandrogens in the this species (Makynen et al , 2000). In zebrafish, the results are more conclusive albeit somewhat negative. During the validation of the fish screen all three laboratories did not detect any effect in ovarian histopathology whilst two of the three laboratories reported some effects on testes histology-increase of interstitial cells and spermatogonia (OECD 2006b). Studies with other fish The AFSS and other bioassays for antiandrogens 9

species include the work by Kinnberg and Toft (2003), where sexually mature male guppies were exposed to a number of oestrogenic and antiandrogenic compounds. Although FL, p,p’ -DDE and oestrogens blocked spermatogonial mitosis, VZ did not have any adverse effects.

2.1.6 Summary of generic effects as markers for antiandrogens Taking everything into account it appears that antiandrogens are not easily detected by the endpoints typically measured in fish short-term tests. Even a combination of more than one endpoint is not always a sensitive means of assigning antiandrogenic activity to a chemical. This was clearly demonstrated in the validation work where FL exposure caused no clear and reproducible alteration in any of the core species in any of the following endpoints: VTG, secondary sex characteristics, spawning status and gonad histology (OECD, 2006b). A follow-up study with FL involving two laboratories produced also variable results; one laboratory reported changes in fecundity but the other did not detect any significant effects on this endpoint (Battelle, 2006, table 1, section 5). The lack of diagnostic marker for antiandrogens in fish has initiated a significant research activity in trying to discover other diagnostic means; gene expression changes and construction of molecular networks that can inform and predict adverse outcomes using chemical-induced changes in mRNA have been employed in this field. This is particularly true for the ZEB where the inability of apical endpoints to detect androgen antagonism was conclusive but also because a plethora of molecular resources are available (e.g. Smolinsky et al , 2010; Wang et al , 2010; Martinovic-Weigelt et al , 2011). Equally, gene expression analysis has been used to detect effects in FHM after exposure to FL (Filby et al , 2007; Wang et al , 2010; Perkins et al , 2011) and VZ (Villeneuve et al , 2007; Ankley et al , 2009) although a substantial effort was also placed in exploiting the secondary sexual characteristics (SSC) present in this species as a more specific endpoint for androgenic xenobiotics (see section 2.2 and table 1, section 5). Although it is anticipated that in the future gene expression analysis will be used more widely in the setting of chemical testing, currently there is a long distance between research and regulatory testing. In this respect we think it is outside the scope of this report to expand further on studies of this nature.

The AFSS and other bioassays for antiandrogens 10

2.2 Species-specific androgen-regulated traits in fish Many male fish display strong sexual dimorphism and offer a range of secondary sexual characteristics (SSC) as potential endpoints for screening androgenic xenobiotics. These include elongation of the fin ray, increase in skin thickness, kidney hypertrophy, nuptial coloration and development of copulatory organs. From the plethora of the fish SSC only four characters have been proposed as potential (anti)androgen biomarkers in chemical testing.

2.2.1 Spiggin in the stickleback The three-spined stickleback offers a great potential for the assessment of reproductive disturbances caused by androgenic xenobiotics due to the pronounced androgen-dependent male secondary sexual characters that presents during its breeding season. These characters include development of nuptial coloration, kidney hypertrophy, territorial and nest-building behaviour. Of these, kidney hypertrophy is by far the more specific and objectively measured response to the rise of endogenous androgens in male fish reaching sexual maturation (Figure 1). It is well established that the kidney hypertrophies under the control of androgens (van Oordt, 1924; de Ruiter and Mein, 1982; Mayer et al , 1990; Borg et al , 1993) to produce a ‘glue’ protein that is used to build the nest out of algae, plant material, sand and detritus. This glue protein was first characterised by Jakobsson and co-workers (1999) and was given the name spiggin from the name of the stickleback in Swedish, the spigg.

Figure 1: Kidney sections of control female stickleback (left), control breeding male stickleback (middle) and androgenised (MT at 1µg/L) female stickleback. Sections were stained with PAS (periodic acid-Schiff).

Spiggin is assembled in the urinary bladder (Jones et al , 2001) and is deposited on suitable nest material by contractions of the male urinary bladder in a process called gluing. There are at least

The AFSS and other bioassays for antiandrogens 11

five different spiggin genes, from which alternative splicing variants can be derived, giving rise to multiple transcripts (Kawahara and Nishida, 2006; 2007). So far spiggin has been shown to exist in at least 14 isoforms, all of which contain polymerization domains. Under natural conditions it should not be present in the kidneys of female fish (similarly to VTG in that should not be present in the blood of male fish) hence it was developed (Katsiadaki et al , 1999) and validated (Katsiadaki et al , 2002) as a specific biomarker for androgens. The effectiveness of different androgens in stimulating hypertrophy of the stickleback kidney tubules, was studied by injecting castrated male sticklebacks with various steroidal androgens at doses of 0.008-25 µg/g BW/day (or control injections were given) for 3 weeks (Borg et al , 1993). The amount of the androgen needed (µg/g BW) to obtain half of the maximal stimulation of the kidney epithelium height was calculated for a number of steroids. Their relative potency if 11KT is rated as 100 were: 11KT(100)>11β-hydroxyandrostenedione(67)>11-ketoandrostenedione(42)>5α-andro- stane-3,11,17-trione (4.4)> 5α-DHT (2.6)> 5β-DHT (1.6)>T (0.3). It was therefore suggested that 11KT is the physiologically relevant androgen in stimulating the stickleback kidney (Borg et al , 1993). Spiggin is to date the only androgen-induced protein in fish and by far the most exploited biomarker for (anti)androgens in both chemical exposure studies and physiological studies (Katsiadaki et al , 2002a; Hahlbeck et al , 2004; Katsiadaki et al , 2006; Jolly et al , 2006; Andersson et al , 2007; Bjorkblom et al , 2007; Allen et al , 2008; Sanchez et al , 2008; Jolly et al , 2009; Macnab et al , 2009; Rushbrook et al , 2010) similar to VTG, a widely used biomarker for oestrogens. We have shown that exposure to a 10% dilution of pulp mill effluent also induces spiggin in female fish (Katsiadaki et al , 2002b). Research at Cefas (Centre for Environment, Fisheries and Aquaculture Science), UK has firmly established that female stickleback kidneys can produce spiggin in response to several model androgens such as methyltestosterone (MT), DHT, T, 11KT, and trenbolone (TB) added to the ambient water. The relative potency of androgens tested in this system on spiggin induction is: MT>DHT>TB>11KT>T (Katsiadaki et al , 2007). The stability of the test chemical in the water and its relative binding affinity for the sex hormone binding globulin are factors (other than receptor affinity) that affect waterborne exposure responses and create this apparent discrepancy on androgen potency between waterborne exposures and internal administration. There are multiple methods by which spiggin can be estimated or measured in the stickleback kidney. The first published method (Borg et al , 1993) for obtaining an estimate of spiggin production in the stickleback kidney employed histological measurements of the kidney secondary proximal epithelium height (KEH). The second published method was an enzyme linked immunosorbent assay (ELISA), which was developed (Katsiadaki et al , 1999) and validated (Katsiadaki et al , 2002) at Cefas.

The AFSS and other bioassays for antiandrogens 12

More recently Sanchez et al , (2008) described an ELISA for spiggin using an antibody against a peptide sequence of spiggin. In addition, spiggin mRNA levels can be detected in the stickleback kidney by real time PCR (Nagae et al , 2007; Hogan et al , 2009). Spiggin protein can also be measured successfully in vitro (Jolly et al , 2006; Jolly et al , 2009; Björkblom et al , 2007; Björkblom et al , 2009) employing a primary kidney cell culture. In the stickleback kidney, previous studies have shown that no specific binding of 11KT or T was detected in either cytosolic or nuclear fractions, although displacement of tritiated 11KT with unlabelled 11KT was observed in the kidney membrane fraction (Jakobsson et al , 1996). Two androgens receptors, ARα and ARβ (AAO83572/3) have been isolated from the stickleback kidney but recent unpublished as yet in situ hybridisation studies revealed that only the ARα is essential for spiggin synthesis (Nagae M., 2010).

2.2.2 Gonopodium in mosquitofish and guppy One of the clearest examples of androgenicity in the aquatic environment has been discovered using morphological changes in female mosquitofish ( Gambusia sp .), living downstream of pulp mill effluent discharges. These fish develop anal fin appendages (gonopodia) that are normally only found in males (Figure 2; Howell et al , 1980; Denton et al , 1985; Howell and Denton, 1989). The causative agent of masculinisation contained within pulp mill effluents displays similar, though not exactly identical, chromatographic properties to androstenedione (Jenkins et al , 2001; Durhan et al , 2002).

Figure 2 (left): Female mosquitofish exposed to pulp and paper mill effluent are masculinised and display an enlarged male-like anal fin Figure 3 (bottom): Male guppy displaying an enlarged anal fin (arrow) Photographs taken from Google images

The AFSS and other bioassays for antiandrogens 13

The response of females to exogenous androgens by means of gonopodium development has been reported very early in time (Turner, 1941; 1942). More recently, Angus et al , (2001) has experimentally induced gonopodia in 11KT-treated female mosquitofish and developed an image analysis system proposing that the mosquitofish anal fin transformation is a useful bioassay system for quantifying the effects of environmental androgens. More recently it was reported that exposure to androstenedione via the water (but not the diet) caused masculinisation of adult female mosquitofish (Stanko and Angus, 2007). To our knowledge there is only one laboratory study that employed this trait in male mosquitofish to detect antiandrogens. Exposure to spironolactone however, an aldosterone antagonist used as a diuretic, with antiandrogenic effects in humans, indicated that this chemical, has androgenic activity in male mosquitofish (Raut et al , 2011). Similar to the mosquitofish, the male guppy ( Poecilia reticulate ) possess a modified tubular anal fin (gonopodium), located directly behind the ventral fin which is flexed forward and used to deliver sperm in the female (figure 3). In addition to the distinct morphology between male and female anal fin, the colour of the tail is also different between males and female and has been used to detect androgenic effluents (Larsson et al , 2002). Bayley et al, (2002) exposed juvenile guppies to FL, VZ and p,p’ -DDE via the food from birth to adulthood and concluded that all three chemicals (in the case of VZ this was not in a dose response manner however) had a clear demasculinising effect by means of reducing gonopodium length. Kristensesn et al , (2006) however failed to reproduce this effect using p,p’-DDE besides the fact that the exposure was longer and both route and chemical concentration were identical to the Bayley et al , study (2002).

2.2.3 Tubercles in the fathead minnow The induction of nuptial tubercles (normally present in sexually mature males) in female FHM has been extensively studied as a biomarker of androgen exposure. The first studies utilise MT as a model androgen; The concentrations of MT used to induce the formation of these tubercles, range from 20–2000μg/L (Smith, 1974; Ankley et al , 2001; Hornumg et al , 2004), which is 40–4000 times the amount required to induce spiggin production in female sticklebacks (Katsiadaki et al , 2006). More recent studies employed TB a masculinising agent; Trenbolone acetate is an anabolic steroid used as a growth promoter in beef cattle in the US and Canada. It is hydrolysed to 17 β-trenbolone, which has been proved a potent environmental androgen both in vitro and in vivo (Wilson et al , 2002). The literature data on TB-induced secondary sexual characters in fathead minnow are very variable. In two studies TB was reported as a very potent androgen (far more than the MT) in The AFSS and other bioassays for antiandrogens 14

inducing tubercle in female FHM at the lowest concentration tested, 50ng/L (Ankley et al , 2003) and 90ng/L (Jensen et al , 2006). Another USEPA study on TB reported a LOEC of 600ng/L (Battelle, 2003). However, three further studies (Seki et al , 2006; Lab 1 and Lab 2, OECD 2006a) failed to observe significant tubercle induction by TB in female FHMs even at the highest concentration tested (5000ng/L). It was suggested that this difference was due to the lack of familiarity of the Japanese laboratories participating this exercise and although this may be very true, it is also important to note that this character is not always unambiguously detected suggesting some objectivity in the measurement.

Figure 4: Typical female (top) and sexually mature male (bottom) fathead minnow (photo courtesy of T, Runnalls, Brunel University, UK)

In contrast, induction of papillary processes on the anal fin of female medaka (see section 2.2.4.) was clearly observed already after 14 days (and 21 days) of exposure to TB and statistical significance was detected at 500ng/L in all three laboratories (OECD 2006a). In the stickleback, TB exposure resulted in statistically significant induction of spiggin at 5000ng/L by all three laboratories during an inter-calibration exercise regardless of the experience the laboratories had with this species (Allen et al , 2008). These results indicate that papillary processes of medaka and spiggin induction in the stickleback are more reproducible endpoints for androgen detection in comparison to the induction of nuptial tubercles in the female FHM where prior knowledge and experience with this species is critical for successful detection. In the past eight years a substantial literature has been generated employing this SSC (reduction in the number of nuptial tubercles in male fathead minnow) to detect antiandrogenic chemicals (see table 1, section 5). In fact this endpoint was refined to include not only number but also intensity of The AFSS and other bioassays for antiandrogens 15

tubercles providing an overall tubercle score. This is calculated by assigning each tubercle a value of 1–3 (Jensen et al , 2001). A value of 1 indicates that a tubercle was present with a height approximately equivalent to the radius, 2 indicates a tubercle was enlarged with an asterisk-like appearance, and 3 indicates a pronounced large rounded tubercle. The antiandrogenic effect of FL in male fathead minnows has been first reported by Panter et al , (2003). However, a FL concentration of 1000μg/L was required to elicit this response, which is 2 times greater than this required to totally inhibit spiggin production in the kidney of photo- periodically stimulated male sticklebacks (Sebire et al , 2008) and 4 times greater than that used to totally block spiggin induction by DHT in non breeding male sticklebacks (Katsiadaki et al , 2006). Similarly, a concentration of 650μg/L failed to produce any effect in this SSC in the male FHM (Battelle, 2003). Two further studies on FL in the FHM were negative; FL exposure at 50 and 500µg/L did not alter any external male secondary sex characters, including tubercle number (Jensen et al , 2004) neither FL exposure at 320µg/L (Filby et al , 2007). In addition, the effect of FL on male tubercle score was variable in the fish screen validation exercise with only one laboratory reporting a reduction in the highest concentration (1000μg/L) tested (OECD 2006b). In a follow-up study, organised by the USEPA, exposure of FHM to FL resulted in a significant reduction of the male tubercle score (Lab 2, Battelle 2006) at 500μg/L (measured 388μg/L) whilst this endpoint was not affected at 1000μg/L (measured 690μg/L; Lab 1, Battelle, 2006). With the exception of one laboratory (Lab 2, Battelle, 2006) the outcome of all the studies is consistent and suggests that the LOEC for FL-induced reduction of tubercle number in male FHM is ≥1000µg/L, a substantially high concentration indicating that the sensitivity of this endpoint is low. Vinclozolin has been tested in the fathead minnow at least 6 times (table 1, section 5). The two studies described in Makynen et al , (2000) did not employ SSC as endpoint. From the remaining four exposures, three were part of a ring test (Biever et al , 2007) and one was independent (Martinovic et al , 2008). All four VZ studies are in agreement in that VZ exposure had an effect on the tubercle count and score (and dorsal pad index in higher concentrations) in the male FHM. The LOECs (as measured concentrations) for tubercle score (the more sensitive SSC) spanned between 150µg/L (Lab B, Biever et al , 2007), 255µg/L (Martinovic et al , 2008), 760µg/L (Lab C, Biever et al , 2007) and 830µg/L (Lab A, Biever et al , 2007) precluding any firm establishment of a NOEC and LOEC for this pesticide. Nevertheless, the above data suggest that VZ is a more potent antiandrogen than FL using as an endpoint the reduction of male FHM SSC. This is not in agreement with the mammalian data suggesting profound differences in receptor binding affinity and/or metabolism of these chemicals between the fathead minnow and the mammalian AR.

The AFSS and other bioassays for antiandrogens 16

The active metabolite of DDT, p,p’-DDE has been tested only once in the FHM (Battelle, 2005) but effects on tubercle score in males have not been quantitatively reported, besides the fact that this endpoint was part of the protocol. It is unclear therefore whether this endpoint was not affected or not measured. We are not aware of any studies using LN or FN, well-known mammalian antiandrogens using the FHM assay.

2.2.4 Anal fin (papillary processes) in medaka A strong sexual dimorphism exists in the Japanese medaka; males have longer and morphologically distinct anal fins than females (Figure 5). Administration of various androgens induce the male-like formation of papillary processes on the anal fin rays in the female medaka (Asahina et al , 1989), including MT (Seki et al , 2004); this character presents therefore a significant potential as an androgen endpoint among the core OECD-recommended species for screening EDCs.

Figure 5: Male medaka (left) where elongation of the anal fin and a number of papilliary proceses are visible and female medaka (right). Photo courtesy of N. Tatarazako, National Institute for Environmental Studies, Tsukuba, Japan).

The relative potency of six physiological androgens in inducing this character in female madaka was studied by Asahina et al (1989) by waterborne exposures and was: 11KT>DHT>5α- androstanediol>Testosterone. 5α-androstanedione and dehydroepiandrosterone had no effect even at the highest concentration tested. The LOEC for 11KT and DHT induction was 10µg/L whilst for 5α- androstanediol and testosterone induction was 50µg/L (Asahina, et al , 1989). These data are highly comparable with the effectiveness of androgenic steroids in inducing spiggin in castrated male fish as presented in section 2.2.1.

The AFSS and other bioassays for antiandrogens 17

To our knowledge, two synthetic androgens have been used for medaka exposures evaluating papillary processes in females; of these MT was used in a full life cycle test (Seki et al , 2004) so although it induced masculinised phenotype in the females the data are not comparable with a fish screen. Highly relevant data were produced however in a ring test during the validation of the 21- day fish screen for evaluating EDCs where TB was tested (OECD 2006a). Following TB exposure, the number of papillary processes in female medaka increased significantly at both 14 and 21 days in the medium concentration tested (500ng/L). Although androgen-induction of these male SSC in female medaka has been proved as a very sensitive endpoint, inhibition of its expression in male fish after exposure to antiandrogens has not demonstrated great sensitivity. For example, the LOEC for FL during the validation work for the fish screen was not determined by any of the four laboratories even at the highest nominal concentration of 1000µg/L (OECD 2006b). Two further studies on FL effects on medaka (Chikae et al , 2004; Kang et al , 2006) did not measure this endpoint so we can’t provide an assessment. The only other medaka exposure to an antiandrogenic chemical reported in the literature used o,p’ -DDE (Papoulias et al , 2003). In this study this endpoint was severely disrupted at low concentrations of the chemical but the design of the study is not comparable to a fish screen; indeed the chemical was injected in medaka embryos (0.005ng/embryo) and the absence of papilliary processes in male fish was noted two months later. All other exposures to antiandrogens did not employ or reported changes in the papillary processes in male medaka such as exposure to VZ and CA (Kiparissis et al , 2003) and to p,p’-DDE (Cheek et al , 2001; Zhang and Hu 2008).

3 The AFSS and other bioassays for antiandrogens.

3.1 The principle of the AFSS The AFSS is a product of long-term research that initiated in the UK. It is utilising the androgen- regulated protein spiggin for assessing androgen modulating activity of chemicals. This bioassay represents the adaptation of the established stickleback androgen bioassay, in which the androgen specific biomarker, spiggin, is measured using an ELISA (Katsiadaki et al , 2002a). In the AFSS, spiggin is induced in female fish by a model androgen (MT or DHT) and exposure to the test compound in a range of concentrations takes place at the same time. The expectation is that chemicals with androgen agonistic activity will induce and enhance spiggin and chemicals with androgen

The AFSS and other bioassays for antiandrogens 18

antagonistic activity will inhibit or reduce the induction of spiggin by the androgen in a dose response manner. Several controls (androgen, test compound, water and solvent) serve the interpretation of results. The activity of the test chemical can therefore be assessed by the induction of spiggin (agonist response) or by blocking (antagonist response) spiggin induction by the model androgen. The bioassay has been finely tuned to provide statistically significant stimulation in the female kidney by a model androgen (MT at 0.5µg/L or DHT at 5µg/L) without concealing the antiandrogenic activity of the test chemical. The maximal spiggin response in female fish after a 21day waterborne exposure has been determined for both androgens at an early stage of the assay development and the AFSS androgen concentration has been selected to give approximately 10% of the maximal MT response and 25% of the maximal DHT response. The AFSS has been used to date in more than 30 in vivo experiments (including the validation work) and has successfully detected all seven mammalian antiandrogens tested (see table 2, section 5). It has demonstrated a modest variability in the LOECs between studies but more than 90% of this variation can be explained by fluctuations in the chemical concentrations. Juvenile male sticklebacks can also be used for the in vivo detection of antiandrogenic activity; female fish, however, provide more consistent results due to the absence of endogenous androgens. Prepubertal male sticklebacks can be used without the need of androgenisation by MT or DHT, only under extremely controlled photoperiodic conditions prior to and during the experiment achieving a fully synchronised reproductive status between the treatment groups (Sebire et al , 2008; Sebire et al , 2009). The endogenous levels of androgens in actively breeding males are extremely high (Mayer et al , 1990) thus, masking the antiandrogenic effect of the test compounds. This is the reason for the surgical removal of testes (the main site of androgen production) in the Hershberger assay.

3.2 Alternative androgenised female fish for detecting antiandrogens Table 2 (section 5) provides a detailed account of fish studies with a similar design to the AFSS (where female fish are stimulated by exogenous androgen administration simultaneously with the putative antiandrogen). Only two fish species (other than the stickleback) have employed this system, the fathead minnow and the mosquitofish. Flutamide successfully antagonised tubercle induction by MT (used at 500ng/L) at 400µg/L (the NOEC was not determined as FL was used at a single concentration) in female FHM after two weeks of exposure (Angley et al , 2004). In two further studies with similar design, tubercle induction by TB (used at 500ng/L) was reduced by VZ at 200µg/L (Martinovic et al , 2008) and by CA at 200µg/L (Ankley et al , 2010). The same effect (reduction in tubercle score in TB-stimulated females) was

The AFSS and other bioassays for antiandrogens 19

observed by exposure to ethinyl-oestradiol (EE2) at only 10ng/L and Bisphenol A (BPA) at 100µg/L. We have also reported the in vivo antiandrogenic activity of oestrogens using the AFSS (Katsiadaki et al , 2006) and subsequent work revealed that the concentrations of oestrogens needed to antagonise the masculinising effect of DHT are very high, at least 10-100 fold higher than those able to elicit an oestrogenic response by means of VTG induction (own unpublished data). However, the fact that the androgenised female FHM assay responds more dramatically to oestrogen administration in comparison to well-known antiandrogens questions the specificity of this bioassay. In addition, in the AFSS the LOECs for both FL and VZ were a fraction of those determined in the FHM studies (table 2, section 5). It is true however that the above studies utilised only one or two concentrations of the antiandrogen, so further testing of this assay is needed to accurately determine its sensitivity. A further study with the fathead minnow using a combined exposure to TB (500ng/L) and FL (500µg/L) was conducted recently but the exposure lasted only for 48hours and the endpoints employed included VTG (no effect), plasma oestradiol (no effect), gene expression changes (Garcia-Reyero et al , 2009) and protein expression changes (Martyniuc et al , 2009). The single relevant study in western mosquitofish employed an androgen (ethynyl-testosterone) to stimulate gonopodium histogenesis in fry and an antiandrogen (FL) to block this effect, unravelling the molecular basis of anal fin development. Although this study was conducted in a very elegant way and elucidated the mechanistic basis of this particular male sexual character, its design is not suited for chemical screening and further refinement and testing is needed before is assessed as a candidate bioassay.

3.3 The principle of the Hershberger bioassay The Hershberger bioassay is a short-term screening test that originated in the 1930’s using accessory tissues of the male reproductive tract and was modified in the 40’s to include androgen- responsive muscles of the male reproductive tract. In the 60’s it was first validated as a screening tool for androgenic and antiandrogenic chemicals (Dorfman, 1969a and 1969b). The bioassay evaluates the ability of a chemical to elicit biological activities consistent with androgen agonists, antagonists, or 5α-reductase inhibitors for a single endocrine class, namely (anti)androgens and by a single mechanism, mediated only by the androgen receptor (AR). This ability is based on the changes in weight of five androgen-dependent tissues, namely the Ventral prostate (VP), seminal vesicles with coagulating glands (SVCG), levator ani-bulbocavernosus (LABC) muscle, paired Cowper’s glands (COWS), and the glans penis (GLANS). The primary model for the Hershberger bioassay is the surgically castrated, peri-pubertal or adult male rat (Ashby and Lefevre, 2000). Accessory sex tissues and glands depend upon androgen

The AFSS and other bioassays for antiandrogens 20

stimulation to gain and maintain weight during and after puberty. When endogenous sources of androgen are low (as a result of castration), the biological activity of exogenous substances can be assessed by the increase (agonist response) in the weights of these accessory sex tissues or by blocking (antagonist response) the activity of administered androgens and by preventing an increase in the weights of these accessory sex tissues. The AFSS is based on exactly the same principle as the Hershberger assay with the following differences: a) female fish are used instead of castrated male rats (both low in endogenous androgens) b) only one tissue is relevant (the kidney versus the five androgens-sensitive tissues in rats), c) chemical administration is via the water rather than by injection and/or gavage, d) the duration is chemical exposure is 21 days as opposed to 10 days in rats. Although it is often stated that the Hershberger bioassay is a relatively rapid and easy to conduct screen the prerequisite for surgical castration is a disadvantage. Surgical removal of testes has serious animal welfare concerns and in addition requires high expertise, costs and labour, factors that have led in the development of several alternative tests. These include the weanling male rat assay (Ashby and Lefevre, 1997); the intact young male rat assay (Cook et al , 1993; O’Connor et al , 1999); the androgen-stimulated immature intact male rat assay (Ashby et al , 2002); and the use of Gonadotrophin Release Hormone-inhibited rats (Ashby et al , 2001; Nellemann et al , 2003). The intact stimulated weanling male rat assay was sought as the strongest alternative model to the castrated rat Hershberger assay and has undergone validation (OECD 2009a); however the results derived from the validation work were not consistent in detecting effects from weak antiandrogens so currently the weanling male rat assay is a guidance document (GD 115, OECD 2009b). Since the Hershberger assay is relatively specific and limited to AR-dependent actions, the Hershberger needs to be complimented by other assays, e.g., inhibition of steroidogenesis to obtain a complete picture of the potential of a chemical to affect the androgen hormone system. The same holds true for the AFSS where in addition 5α-reductase inhibitors are not detected (testosterone is a weak androgen in fish).

3.4 Antiandrogen testing in the AFFS and the Hershberger assays Between 2001 and 2007, the rat Hershberger assay has undergone an extensive validation programme including a series of intra and inter-laboratory studies to demonstrate the reliability and reproducibility of the assay. These validation studies were conducted with a potent reference androgen (testosterone propionate; TP), two potent synthetic androgens (trenbolone acetate, TB and MT), a potent antiandrogenic pharmaceutical (FL), several weakly antiandrogen ic pesticides (LN, vinclozolin; VZ, procymidone; PRO, p,p’ -DDE), a potent 5α-reductase inhibitor (finasteride; FIN) and

The AFSS and other bioassays for antiandrogens 21

two known negative chemicals (dinitrophenol; DNP and nonylphenol; NP) (OECD, 2006c; Owens et al , 2006; OECD, 2007, OECD 2008). The list of antiandrogens tested in the AFSS is conveniently almost identical (OECD 2010) to the list of chemicals tested in the validation of the Hershberger. However, due to the different routes of administration of test chemicals between the two assays it was proved a real challenge to produce a basis for comparing the sensitivities and specificities of the responses to AR antagonists. The LOECs and NOECs are expressed in mg/kg/day in the Hershberger assay and in µg/L in the AFSS. One suggestion was to estimate the total amount of chemical administrated in each animal (fish or rat) over the duration of the exposure. This is relatively easily done for the Hershberger assay but very difficult to conduct in the AFSS where exposure is waterborne. Information such as diffusion (flux) coefficient(s) across gills and epithelial layers, affinity for the steroid binding globulin in fish plasma, excretion rates and bioaccumulation rates are needed in order to estimate the amount of chemical that is bioavailable to fish; these data are clearly not available for each (if any) of the tested chemicals. A calculation of the theoretical maximum amount of chemical that an individual fish is exposed to can be achieved if the flow rate, the exposure concentration, the duration of the exposure and the number of fish in the aquarium are known. Similarly in the Hershberger assay the amount of chemical administered in each rat can be calculated by multiplying the daily dose by the duration of exposure. In the AFSS, DHT is used at a nominal concentration of 5µg/L, provided at a flow rate of 100ml/min over 21 days in each aquarium containing 20 fish (this was the design of the AFSS validation data). This means that the total amount of chemical potentially taken by a single fish over 21 days is 0.756mg DHT (if MT is used then the amount is 0.0756mg MT). In the Hershberger assay TP at 0.2mg/kg/day or 0.4mg/kg/day is injected daily into each animal over 10 days, resulting in the total amount of TP into each animal of 0.6mg or 1.2mg respectively (assuming an average weight of 300g for each rat). These figures suggest that androgen dose used to stimulate responses in the AFSS and the Hershberger assay when expressed as total chemical theoretically available to each animal are relatively comparable (although MT levels in the AFSS are much lower than TP in the Hershberger). The assumption that all the test chemical passing through an aquarium is fully absorbed results in a significant overestimation of the chemical available to fish since measured concentrations of the test compounds in the aquaria water were never undetectable (typically between 60 and 80% of the nominal concentrations) in the AFSS validation work, suggesting that some sort of equilibrium is reached between the chemical delivery rate and the absorption rate.

The AFSS and other bioassays for antiandrogens 22

Assuming however that a stable equilibrium is reached between the chemical in the water and the chemical inside the fish is equally incorrect. It has been demonstrated that static exposure of sticklebacks to steroids (E2 and T) resulted in a rapid bioconcentration in their plasma and were up to 50-fold (E2) and 200-fold (T) greater than the actual levels of steroid measured in the exposure water, within the first 6 hours of exposure (Maunder et al , 2007). Although this comparison is of limited value, one could argue that overall the spiggin stimulation in the stickleback kidney is achieved at similar if not lower androgen levels compared to the Hershberger assay besides the longer duration of exposure and the constant flow of the chemical. For environmental antiandrogens, the amount of total chemical delivered through the aquaria is far less than that administered by oral gavage. The LOEC for LN for example in the AFSS was determined as 250µg/L which means a theoretical maximum of 37.8mg was available for each fish. The LOEC for LN in the Hershberger validation work was determined as 100mg/g/day giving a theoretical maximum amount of 300mg for each rat. Table 3 (section 5) presents the LOECs determined in both the AFSS and the Hershberger for seven chemicals for which data were available. The majority of these data come from the validation work for both assays; only few are supplemented by published or unpublished studies for comparative purposes. The fold change in the endpoints measured in each assay between the androgen stimulated control and the test compound highest concentration tested is also provided in table 3 (section 5) in order to compare the dynamic range of each assay. The relative potency of antiandrogens in inhibiting spiggin induction by DHT (5 µg/L) in the so far conducted AFSS tests is Flutamide>Fenitrothion>Vinclozolin> Procymidon>Linuron≥ DDE>NP. In the Hershberger this order is Flutamide>>Vinclozolin>Procymidon>Fenitrothion>DDE>Linuron>NP (Table 3, section 5). The similarity between these datasets not only in terms of detection but also ranking is remarkable, considering the reported differences between mammalian and fish responses to EDCs. The only significant difference between the responses in the AFSS and the Hershberger is that FL is a far more potent antiandrogen in the Hershberger (there is almost 100-fold difference between the LOEC for FL and the next more potent environmental antiandrogen) than in the AFSS (where FN is almost as potent as FL and the LOEC for VZ is only 2-fold higher than for FL). Flutamide is a pharmaceutical specifically designed to target the human AR hence it is not a surprise that its affinity for the rat AR is higher to this for the stickleback ARs. In fish studies (other than the AFSS) FL has been characterised as a particularly weak antiandrogen (see tables 1 and 2, section 5). In the AFSS however, FL is to date the most strong androgen antagonist, with a potency that only FN has superseded in a single experiment (OECD 2010). Besides, in the first flow-through experiment with

The AFSS and other bioassays for antiandrogens 23

FL in the AFSS a statistically significant reduction in spiggin was observed when FL was used at only 10µg/L (Katsiadaki et al , 2006). In relation to other environmental antiandrogens tested, the AFSS has shown a high sensitivity with LOECs for VZ, FN, LN, PRO and DDE relatively close to this for FL (table 3, section 5). This observation is potentially of great importance as it may suggest that the ligand binding domain in the stickleback ARα (this is essential for expression of spiggin in the kidney) may react with environmental antiandrogenic chemicals more widely and at lower concentrations than the rat AR; in the context of regulatory testing this may provide some sort of additional assurance for detecting weak androgen antagonists. The weak oestrogen nonylphenol (NP) was utilised as a negative chemical in the validation of the Hershberger bioassay. In our opinion this is somehow erroneous as oestrogens have a relatively high affinity for the AR, so they may act as pure antiandrogens by occupying the receptor but not activating it (Sohoni and Sumpter, 1998). Besides, diethylstibesterol, oestradiol and NP have been tested in the Hershberger assay and antagonised the androgen action in at least one sensitive tissue (Yamazaki et al , 2003; 2004; Tyl et al , 2006). In addition, in the Hershberger validation work, although exposure to NP produced an overall negative result, it gave a small but significant organ weight change (14%) indicating a possible antiandrogenic effect (OECD 2007). The concentration at which NP was tested as a negative reference substance was based on the uterotrophic dose, so higher concentrations could possibly demonstrate a more pronounced antiandrogenic effect. In the AFFS we have shown that in vivo exposure to high concentrations of oestrogens (i.e. oestradiol and ethinyloestradiol at ≥100ng/L, oestrone at ≥ 1000ng/L and NP≥500µg/L) also inhibit androgen-induced spiggin synthesis (Katsiadaki et al , 2006; own unpublished data). This antagonism was also observed in vitro using a stickleback kidney primary culture (Jolly et al , 2009) but at even higher concentrations in relation to pure antiandrogens. Both findings are in line with previous observations by Oguro (1957; 1958), who first reported that oestrogens do not stimulate kidney hypertrophy in the stickleback but do result in regression of the kidney hypertrophy. It is possible that this effect is enhanced by another mechanism, not directly related to the affinity of oestrogens for the AR. Alternative mechanisms for the antiandrogenic effect of oestrogens could include non- genomic membrane receptors, feedback control of sex steroid levels to gonadotrophins or suppression of AR expression as demonstrated in the stickleback (Olsson et al ., 2005) where AR mRNA levels were downregulated in castrated males that received oestradiol implants. There are two significant advantages in the AFSS in comparison to the Hershberger; the high resolution that offers (see fold change in table 3, section 5) and the use of intact (not surgically

The AFSS and other bioassays for antiandrogens 24

castrated animals). On the other hand, the Hershberger assay has a long track record of use in both industry and academia and is a highly relevant mammalian system.

4 Overall discussion and conclusions

The AFSS has been extensively used for the detecting on (anti)androgenic chemicals (table 2, section 5) and provided the first evidence of the antiandrogenic activity of Linuron and Fenitrothion in fish (Katsiadaki et al , 2006). In addition to the high specificity and sensitivity of spiggin as an endpoint, the AFSS assay has a huge dynamic range spanning from less than 100 spiggin units in a non-stimulated female kidney to over 50,000 and well over 150,000 spiggin units in DHT and MT stimulated female kidneys respectively. The tubercle number for example in a fathead minnow is 0 in a control female and 10-12 in a fully androgen-stimulated female (Ankley et al , 2004). Even if the tubercle score is employed where each tubercle may be theoretically multiplied by a maximum factor of 3, the resulting range is 0-32, much lower than the range available in the AFSS. The high resolution of spiggin as an endpoint provides a firm basis for unequivocal detection of chemicals with both androgenic and antiandrogenic activity in a statistically robust and dose related fashion. The AFSS has demonstrated an excellent agreement with the Hershberger assay, the ‘golden standard’ test for the detection of androgens and antiandrogens in mammals, indicating that it may not be relevant only for fish but also for mammalian research on this increasingly worrying class of EDCs. The AFSS presents also a much higher resolution to the Hershberger assay (see table 3, section 5) and a greater sensitivity for environmental antiandrogens as judged from the proximity of the LOECs between the pharmaceutical analogue FL and five pesticides. However, direct comparison of the sensitivities of the two assays is not possible due to the complexity of determining the bioavailable chemical concentration. Nevertheless, the close proximity in ranking the potency of seven compounds merits further research on the possibility of the AFSS to be used as an alternative to the Hershberger for chemicals that demonstrate androgen antagonistic activity in vitro and where exposure of aquatic biota is a major concern. At this point of time the AFSS is by far the more sensitive and reliable screen for antiandrogenic chemicals in fish. The presence of two ARs in the stickleback unlike cyprinids (including the zebrafish and the fathead minnow) may be linked with its high sensitivity in detecting this class of EDCs. Spiggin is the only androgen-induced protein known in fish and is usefulness as a biomarker is directly analogous to this of vitellogenin. The only unfortunate fact is that this unique character

The AFSS and other bioassays for antiandrogens 25

exists only in the stickleback, a fish that besides its large geographical distribution and extensive body of literature on its biology, it has not as yet gained substantial interest by toxicologists. The species certainly presents an ‘inconvenience’ in terms of infrastructure for chemical testing since its upper temperature limit does not exceed 18ºC, almost 10 degrees below the temperatures used for zebrafish, medaka and fathead minnow exposures. In addition to the unique androgen biomarker, however, other advantages of the stickleback as a model species include ability to thrive in fresh and sea water, the easiness to maintain and reproduce in the laboratory, a fully sequenced genome and in the context of EDCs work, the presence of a genetic sex marker.

The AFSS and other bioassays for antiandrogens 26

5 Tables

List of abbreviations

11KT 11-ketotestosterone red. reduced EE2 17 α-ethinyloestradiol SSC secondary sexual characteristics E2 17β-oestradiol SV seminal vesicles plus coagulating glands / and yes significant change(s) AA androgen(s) SPL spironolactone AA antiandrogen(s) Static static renewal BPA bisphenol A T testosterone conc. concentration TP testosterone propionate C control TB trenbolone COWS Cowper’s glands VP ventral prostate CA cyproterone acetate VZ vinclozolin d day VTG vitellogenin dpf days post-fertilisation w week dph days post-hatch wph week post-hatch dec. decreased DDE dichloro-diphenyl-dichlorethylene DHT dihydrotestosterone CARP carp ET ethynyl testosterone FHM fathead minnow F female(s) GUP guppy FN fenitrothion MED medaka Flow flow-through exposure MF mosquitofish FL flutamide MM mummichog GLANS Glans penis SB three-spined stickleback h hour hpf hours post-fertilisation inc. increased Separation lines ind. induced separation between publications or I injection LABC levetor ani and bulbocavernous muscle studies LN linuron LOEC Lowest Observed Effect Concentration separation between chemicals M male(s) MT methyltestosterone separation between species mth month no data no data presented NOEC No Observed Effect Concentration no no significant effect(s) NP nonylphenol nb number PRO procymidone The AFSS and other bioassays for antiandrogens 27

Table 1. Synopsis of in vivo fish exposures to well known antiandrogens.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

FHM FL Flow 21d M/F 60, 650 47, 510 Fecundity 47 47 ↓ SSC > 510 > 510 no nb tubercles & fat pad VTG > 510 > 510 no Plasma E2 47 (M) 510 (M) ↓ no effect in F Plasma 11KT 47 (M) 510 (M) ↑ not detected in F Plasma T 47 510 ↑ Histo 510 (F) 510(F) ↑ inc. atretic follicles; no effect in M FHM FL Flow 14d M/F 60, 650 47, 510 Fecundity 47 510 ↓ nb tubercles & fat pad SSC > 510 > 510 no Battelle 2003 VTG > 510 > 510 no Plasma E2 47 (M) 510 (M) ↓ no effect in F Plasma 11KT/ T > 510 > 510 no Histo 510 (F) 510(F) ↑ inc. atretic follicles; no effect in M no dose-related pattern for tubercle & fat pad; no FHM FL Flow 14d M/F* 60, 350, 650 47, 260 , 510 SSC > 510 (M) > 510 (M) no *non-spawning effect in F VTG 47 (F) 260 (F) ↑ no effect in M Plasma E2/11KT/T > 510 > 510 no Histo > 510 > 510 no 100, 320, 95.3, 320.4, FHM FL Flow 21d M/F SSC 320.4 (M) 938.6 (M) ↓ dec. nb tubercles; no effect in F 1000 938.6 Panter et al. but same increase in WC compared to SC; no 2004 VTG < 95.3 (F) 95.3 (F) ↑ effect in M FHM FL Flow 21d M/F 50, 500 63, 651 Fecundity 63 651 ↓ SSC > 651 > 651 no VTG 63 651 ↑ Jensen et al. Plasm E2/ T 63 (F, T) 651 (F, T) ↑ no effect in M or E2 levels 2004 F: dec. mature oocytes, inc. atretic follicles; M: Histo 63 651 yes spermatocyte degeneration & necrosis

100, 500, Jensen and comparison ELISA kits FHM FL Flow 21d F no data VTG 100 500 ↑ no effect at 1000 µg/L with samples from 1000 Ankley 2006 ring test 100, 500, Lab 7: 68.7, FHM FL Flow 21d M/F Fecundity 354 754 ↓ inhibition of egg production 1000 354, 754 SSC 354 (M) 754 (M) ↓ dec. tubercle score ; none in F VTG > 754 > 754 no dec. median ovarian staging scores; no effect in Histo 68.7 (F) 354 (F) ↓ M OECD 2006b 100, 500, Lab 10: 83.5, FHM FL Flow 21d M/F Fecundity 83.5 875 ↓ inhibition of egg production 1000 445, 875 SSC > 875 > 875 no VTG > 875 > 875 no Histo > 875 > 875 no The AFSS and other bioassays for antiandrogens 28

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

100, 500, Lab 4: 88.8, FHM FL Flow 21d M/F Fecundity 464 940 ↓ inhibition of egg production 1000 464, 940 SSC > 940 > 940 no VTG > 940 > 940 no in F significant only at 464 µg/L Histo 88.8 (M) 464 (M) ↑ inc.spermatogonia OECD 2006b dec. spermatocytes & in mineralization in the 464 (M) 940 (M) ↓ collecting duct 464 (M) 940 (M) ↑ inc. median testicular staging scores < 88.8 (F) 88.8 (F) ↑ inc. atresia of immature oocyte 100, 500, Lab1: 74, FHM FL Flow 21d M/F Fecundity > 690 > 690 no dec. nb eggs 1000 340, 690 Fertility 340 690 ↓ dec. fertile eggs SSC (M) > 690 > 690 no tubercle score 340 690 ↓ dec. fat pad index VTG > 690 > 690 no Histo > 690 > 690 no Battelle 2006 100, 500, Lab 2: 89, FHM FL Flow 21d M/F Fecundity 388 822 ↓ dec. nb eggs 1000 388, 822 Fertility 388 833 ↓ dec. fertile eggs SSC (M) 89 388 ↓ dec. tubercle score & count 388 822 ↓ dec. fat pad index & score VTG > 822 > 822 no Histo 89 388 F: inc. atresia of oocyte; M: inc. median testicular 412 µg FL FHM FL Flow 21d M/F 320 SSC (M) > 412 > 412 no tubercles count & grade eq/L > 412 > 412 no fat pad index & grade VTG < 412 (F) 412 (F) ↑ no effect in M Filby et al. 2007 mRNA in liver and Erβ, ERγ (M, ind.), AR (M. red.), vasa, (M, red.), < 412 412 yes* amh (M, red.),dmrt1 (M, red.), sf1 (F, red.), GHR *depending on tissue gonads (red.), IGF(red.), CYP19 (M, ind.), 11β-HSD (ind.) 4 hpf to 75, 150, 300, 89, 162, 267, FHM VZ Flow 34 & 90 M/F Fecundity/Fertility > 1170 > 1170 no 600, 1200 543, 1170 dph Sex ratio > 1170 > 1170 no Makynen et al. 175.8 (M, FHM VZ Flow 21d M/F 200, 700 175.8, 706 Plasma E2/ T 706 (M, E2) ↑ no effect in F or T levels 2000 E2) reduction in oocyte size & retarded oocyte Histo 175.8 (F) 706 (F) yes maturation (qualitative assessment); no effect in M

The AFSS and other bioassays for antiandrogens 29

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L) FHM VZ Flow 21d M/F 100, 400, 700 60, 255, 450 Fecundity < 60 60 ↓ dec. nb eggs SSC 60 (M) 255 (M) ↓ dec. tubercule score; no effect in F 255 (M) 450 (M) ↓ dec. dorsal pad index; no effect in F VTG 60 (F) 255 (F) ↑ no effect in M Plasma T* > 450 > 450 no *no analysis on ex vivo E2 > 450 > 450 no only measured in F plasma E2 due to ex vivo KT (M) 60 255 ↑ contamination ex vivo T < 60 (F) 60 (F) ↑ no effect in M Martinovic et Histo 255 (F) 450 (F) ↑ inc. severity of atresia; no effect in M al. 2008 inc. AR and 11β-HSD transcript levels in gonads & mRNA 255 (M) 450 (M) ↑ AR transcript levels in dorsal pads; no effect in F FHM VZ Flow 13d M/F 200, 700 158, 590 SSC 158 (M) 590 (M) ↓ dec. tubercule score; no effect in F VTG > 590 > 590 no Plasma E2/T > 590 > 590 no inc. AR transcript levels in gonads; not reported mRNA 158 (M) 590 (M) ↑ in F Lab A: 75, FHM VZ Flow 21d M/F 100, 300, 900 Fecundity/Fertility 280 830 ↓ dec. nb eggs, fertile eggs & spawns/female/day 280, 830 dec. tubercule score & count; no effect on fat pad SSC 280 (M) 830 (M) ↓ weight, score and index; not reported in F VTG > 830 > 830 no Plasma E2/T > 830 > 830 no inc. proportion of spermatogonia, testicular degeneration & interstitial cells; altered Histo < 75 (M) 75 (M) ↑ proportion of spermatocytes/ spermatids ; no effect in F Lab B: 150, FHM VZ Flow 21d M/F 100, 300, 900 Fecundity/ Fertility 150 370 ↓ dec. nb eggs, fertile eggs & spawns/female/day 370, 1200 Biever 2007 dec. tubercule score, fat pad weight & index; not SSC < 150 (M) 150 (M) ↓ reported in F 150 (M) 370 (M) ↓ dec. tubercule count; not reported in F 370 (M) 1200 (M) ↓ dec. fat pad score; not reported in F VTG > 1200 > 1200 no Plasma E2 150 (F) 370 (F) ↓ not reported in M Plasma T 370 (M) 1200 (M) ↓ not reported in F inc. proportion of spermatogonia, testicular Histo <150 (M) 150 (M) ↑ degeneration & interstitial cells; altered proportion of spermatocytes/ spermatids

inc. oocyte atresia, perifollicular cell 370 (F) 1200 (F) ↑ hyperplasia/hypertrophy & interstitial fibrosis

The AFSS and other bioassays for antiandrogens 30

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

Lab C: 84, FHM VZ Flow 21d M/F 100, 300, 900 Fecundity/ Fertility 760 760 ↓ dec. nb eggs, fertile eggs & spawns/female/day 270, 760 VTG 84 270 ↑ Biever et al. dec. tubercule score; no effect on tubercle count, SSC 270(M) 760 (M) ↓ 2007 fat pad weight, score & index; not reported in F Plasma E2/T > 760 > 760 no Histo > 760 > 760 no p, p'- FHM Flow 21d M/F 0.02, 0.2 0.022, 0.167 Fecundity > 0.167 > 0.167 no DDE VTG > 0.167 > 0.167 no Battelle 2005 plasma E2/11KT/T > 0.167 > 0.167 no 11KT not detected in F Histo 0.022 (F) 0.167 (F) ↑ inc. proportion of atretic follicles; no effect in M

Diet (0.06 0.02, 0.2, 2, Chikae et al. MED FL 7d M/F n/a VTG 0.02 (F) 0.2 (F) ↑ no effect in M g/10 fish) 20 mg/g diet 2004 Nozaka et al. MED FL Flow 21d M/F no data 90.4-1470 VTG > 1470 > 1470 no 2004 93.8, 188, 101, 202, MED FL Flow 21d M/F 375, 750, 397, 787, Fecundity 787 1560 ↓* *significant 1 st & 2 nd week, but not 3 rd week 1500 1560 (although still dec.) Fertility 787 1560 ↓* Kang et al. 2006 VTG > 1560 > 1560 no Histo > 1560 > 1560 no testis-ova present in M but not dose related 100, 500, Lab 3: 95.5, MED FL Flow 21d M/F Fecundity < 100 100 ↓ inhibition of egg production 1000 518, 1060 SSC > 1060 (M) > 1060 (M) no papillary processes; none in F VTG > 1060 > 1060 no Histo 518 (M) 1060 (M) ↑ inc. median testicular staging scores 518 (F) 1060 (F) ↓ dec. median ovarian staging scores OECD 2006b 100, 500, Lab 5: 94.9, MED FL Flow 21d M/F Fecundity > 880.7 > 880.7 no 1000 434.5, 880.7

SSC > 880.7 (M) > 880.7 (M) no papillary processes; none in F VTG > 1060 > 1060 no Histo > 880.7 > 880.7 no

The AFSS and other bioassays for antiandrogens 31

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

100, 500, Lab 4: 97.4, MED FL Flow 21d M/F Fecundity > 996 > 996 no 1000 501, 996 SSC > 996 (M) > 996 (M) no papillary process; none in F VTG < 100 (M) 100 (M) ↑ in F significant only at 501 µg/L Histo < 97.4 (F) 97.4 (F) ↓ dec. post-ovulatory follicles; no effect in M OECD 2006b 100, 500, Lab 6: 55.8, MED FL Flow 21d M/F Fecundity > 552 > 552 no 1000 221, 552 SSC > 552 (M) > 552 (M) no papillary processes; none in F VTG > 552 > 552 no Histo > 552 > 552 no dec. nb of males with advanced stages of MED VZ Static 1-100dph M/F 2500 no data Histo < 2500 (M) 2500 (M) ↓ spermatogenesis > 2500 (F) > 2500 (F) no Sex ratio > 2500 > 2500 no 1000, 5000 Kiparissis et al. MED VZ Static 1-100dph M/F no data Histo <1000 (M) 1000 (M) ↓ dec. in the density of mature spermatozoa (Ronilan®) 2003 2 M with testis-ova dec. nb of males with advanced stages of 1000 (M) 5000 (M) ↓ 5000 µg/L Ronilan® spermatogenesis 1000 (F) 5000 (F) ↑ elevated nb of atretic oocytes Sex ratio > 5000 > 5000 no MED CA Static 1-100dph M/F 1, 10 no data Histo 1 (M) 10 (M) ↓ dec. density of mature spermatozoa Kiparissis et al. 1 M & 2 M with testis- 1 (M) 10 (M) ↑ inc. testicular fibrosis ova at 1 µg/L & 10 2003 < 1 1 ↓ dec. nb fish with advanced stages of gonogenesis µg/L Sex ratio > 10 > 10 no 0.005, 0.05, o, p'- no anal fin breeding tubercles or processes MED I embryos* M/F 0.5 n/a SSC 0.005 0.05 ↓ DDE except 1 fish at 0.005 ng/ embryo Papoulias et al. *measurement on ng/embryo 2003 adults (~ 2months) few vitellogenic oocytes & inc. oocyte atresia; no Histo < 0.005 (F) 0.005 (F) yes effect in M p, p'- MED Flow 2mth M 1, 5, 20, 100 no data Intersex 20 100 yes 25% occurence (0% in control) DDE VTG-1 and VTG-2 up-regulated with dose- Zhang and Hu VTG mRNA < 1 1 ↑* response relationships 2008 *assumed LOEC from CHG-H and ERα up-regulated; CHG-L, U-shape graph other mRNA < 1 1 ↑* relationship

The AFSS and other bioassays for antiandrogens 32

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L) p, p'- 0dph- 0.5, 1, 2.5, 0.2, 0.5, 1.4, MED Flow M/F Fecundity 1.4 4.3 ↓ nb eggs in adults DDT 2wph* 7.5 4.3 *transferred in clean Fertility < 0.2 (F) 0.2 (F) ↓ water up to 8 and 0.2 (M) 0.5 (M) ↓ 17wph VTG > 4.3 > 4.3 no both in juveniles or adults Cheek et al. female skewed in adults Sex ratio 1.4 4.3 yes 2001 p, p'- 0dph- 0.5, 1, 2.5, 0.3, 0.7, 1.9, MED Flow M/F Fecundity > 5.2 > 5.2 no DDT 8wph* 7.5 5.2 *transferred in clean Fertility < 0.3 0.3 ↓ water up to 13wph VTG > 4.3 > 4.3 no both in juveniles and adults Sex ratio 0.7 1.9 yes female skewed in adults 53% of ZEB FL Flow 21d M/F 10, 100, 1000 nominal Fecundity 100 1000 ↓ number of clutches and eggs within 5d VTG > 1000 > 1000 no inc. interstitial cells, Sertoli cell hypertrophy, size Wester et al. Histo < 10 (M) 10 (M) yes of early gonocyte, spermatogonia & dec. 2003 spermatocyte cysts; no effect in F 53% of 0hpf- male skewed in juveniles in C from treated *paradoxical ZEB FL Flow F1 10, 100, 1000 nominal Sex ratio > 1000 > 1000 no 42dph parents* masculinisation within 5d 100, 500, Lab 12: 76.6, ZEB FL Flow 21d M/F Fecundity > 788 > 788 no 1000 250, 788 VTG > 788 > 788 no Histo 76.6 (M) 250 (M) ↑ inc. interstitial cells 250 (M) 788 (M) ↑ inc. spermatogonia 250 (M) 788 (M) yes Sertoli cell hypertrophy 250 (M) 788 (M) ↓ dec. median testicular staging scores > 788 (F) > 788 (F) no 100, 500, Lab 14: 74.7, OECD 2006b ZEB FL Flow 21d M/F Fecundity > 730 > 730 no 1000 397, 730 VTG > 730 > 730 no Histo 74.7 (M) 397 (M) ↑ inc. interstitial cells 397 (M) 730 (M) ↑ inc. spermatogonia > 730 (F) > 730 (F) no 100, 500, Lab 6: 55.8, ZEB FL Flow 21d M/F Fecundity > 552 > 552 no dec. total nb eggs 1000 221, 552 VTG > 552 > 552 no Histo > 552 > 552 no staging scores not available The AFSS and other bioassays for antiandrogens 33

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

1, 10, 100 Diet (40 µ/mg food GUP FL mg/20 30d M n/a Sperm count* < 1 1 ↓ (i.e. 2, 20, fish) 200 µg/fish) Baatrup and *35% mortality at 100 µg/mg food, excluded SSC* > 10 > 10 no coloration Junge 2001 from analysis Courtship behaviour* < 1 1 ↓ dec. nb sigmoid display

1 10 yes affected posturing behavior & sigmoid display juveni 0.01, 0.1 GUP FL Diet to adult n/a Sperm count > 0.1 > 0.1 no les µg/mg food SSC (M) < 0.01 0.01 ↓ gonopodium length Bayley et al. < 0.01 0.01 ↑ inc. blue coloration intensity 2002 Courtship behaviour 0.01 0.1 ↓ dec. nb & duration sigmoid display (M) 1, 10, 100 Diet (40 35% mortality at 100 µg/mg food dec. nb of spermatogenetic cysts & inc. nb Kinnberg and µg/mg food; same GUP FL mg/20 30d M n/a Histo < 1 1 yes (i.e. 2, 20, spermatozeugmata in the ducts Toft 2003 study than Baatrup fish) 200 µg/fish) and Junge 2001 1, 10, 100 Diet (40 µg/mg food GUP VZ mg/20 30d M n/a Sperm count* > 10 > 10 no (i.e. 2, 20, fish) Baatrup and *15% mortality at 100 200 µg/fish) µg/mg food, excluded Junge 2001 Coloration* > 10 > 10 no dec. only significant at 1 µg/mg food from analysis affected posturing behavior & sigmoid display Courtship behaviour* > 10 > 1 0 no but significant only at 1 µg/mg food juveni 0.1, 10 GUP VZ Diet to adult n/a Sperm count < 0.1 0.1 ↓ dec. sperm cell count les µg/mg food dec. gonopodium length but only significant at SSC (M) > 10 > 10 no Bayley et al. 0.1 µg/mg food 2002 < 0.1 0.1 ↑ inc. blue coloration intensity Courtship behaviour < 0.1 0.1 ↓ dec. nb & duration sigmoid display (M) 1, 10, 100 Diet (40 µg/mg food Kinnberg and same study than GUP VZ mg/20 30d M n/a Histo > 100 > 100 no Baatrup and Junge (i.e. 2, 20, Toft 2003 fish) 2001 200 µg/fish) The AFSS and other bioassays for antiandrogens 34

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

virgin 0.1, 1, 10 *average at 1.8, GUP VZ Diet 30d n/a Fecundity > 10 > 10 no first-clutch size F µg/mg food* 18,180 mg/kg bw

0.1, 1, 10 negative effect on size of the first clutch produced GUP VZ Diet 30d M n/a Fecundity 0.1 1 ↓ by unexposed virgin F the exposed M Bayley et al. µg/mg food* inseminated 2003 *average at 1.8, Sperm count 0.1 1 ↓ 18,180 mg/kg bw Courtship behaviour 1 10 ↓ nb sigmoid displays 0.1, 1, 10 Diet (40 p, p'- µg/mg food ↑ & sperm count higher at 0.1 µg/mg food than C, GUP mg/20 30d M n/a Sperm count* < 0.1 0.1 DDE (i.e. 0.2, 2, 20 ↓ but at 1 µg/mg food lower than C fish) µg/fish) Baatrup and * 70% mortality at 10 Coloration* > 1 > 1 no µg/mg food, excluded Junge 2001 from analysis Courtship behaviour* < 0.1 0.1 ↓ dec. nb sigmoid display

0.1 1 yes affected posturing behavior & sigmoid display p, p'- juveni 0.01, 0.1 GUP Diet to adult n/a Sperm count < 0.01 0.01 ↓ dec. sperm cell count DDE les µg/mg food SSC (M) 0.01 0.1 ↓ gonopodium length Bayley et al. 0.01 0.1 ↑ inc. blue coloration intensity 2002 Courtship behaviour > 0.1 > 0.1 no (M) 0.1, 1, 10 Diet (40 70% mortality at 10 p, p'- µg/mg food Kinnberg and µg/mg food; same GUP mg/20 30d M n/a Histo 0.1 1 yes dec. nb spermatogenetic cysts DDE (i.e. 0.2, 2, 20 Toft 2003 study than Baatrup fish) µg/fish) and Junge 2001 0- p, p'- juveni 0.01, 0.1, 1 GUP Diet 208/212d n/a Sex ratio > 0.1 > 0.1 no DDE les µg/mg food ph SSC (M) > 0.1 > 0.1 no gonopodium Kristensen et al. 100% mortality in 1 inc. orange coloration area & index 0.01 0.1 ↑ 2006 µg/mg food Competition > 0.1 > 0.1 no behaviour (M) Courtship behaviour > 0.1 > 0.1 no (M)

The AFSS and other bioassays for antiandrogens 35

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L)

100, 500, 101.2, 517.3, SB FL Flow 21d M/F * VTG > 714.7 > 714.7 no 1000 714.7 Spiggin 101.2 (M) 517.3 (M) ↓ no effect in F Nest building 101.2 517.3 ↓ dec. digging activity Sebire et al. *photoperiodically behaviour (M) 2008 manipulated fish < 101.2 101.2 ↓ dec. nb nest Courtship behaviour < 101.2 101.2 ↓ dec. overall courtship behaviourr (M) < 101.2 101.2 ↓ dec. in nb & duration of zigzags 101.2 517.3 ↓ dec. nb dorsal pricking 0.89, 53.9, SB FN Flow 21d M/F * 1, 50, 200 VTG > 135.5 > 135.5 no 135.5 Spiggin 0.89 53.9 ↓ no effect in F Nest building 0.89 53.9 ↓ dec. nb nest Sebire et al. *photoperiodically behaviour (M) 2009 manipulated fish Courtship behaviour < 0.89 0.89 ↓ dec. in nb and duration of zigzags (M) < 0.89 0.89 ↓ dec. nb dorsal pricking 0.89 53.9 ↓ dec. nb bites towards F 0.01, 0.10 Nest building SB CA I 21d M/F mg/ 20 µl n/a < 0.01 0.01 ↓ delay in the onset of nest building behaviour (M) injection < 0.01 0.01 ↓ dec. total duration of displacement fanning Rouse et al. Agressive behaviour < 0.01 0.01 ↓ dec. nb of charges to another M 1977 (M) Courtship behaviour < 0.01 0.01 ↓ dec. nb zigzags & dec. chasing & biting F (M) dec. secondary spermatocytes & inc. Histo (M) < 0.01 0.01 ↑/↓ spermatozoa 4.17, 41.7, MF SPL Static 35d F no data SSC < 4.17 4.17 ↑ 4:6 anal fin ray elongation ratios * calculated from 10, 104.1, 208.2 Raut et al. 2011 * 100, 250, 500 nM VTG 41.7 104.1 ↓ 2 , 14 , 100, CARP FL Flow 15d M/F no data plasma E2 100 (F) 700 (F) ↓ not reported in M 700 plasma 11KT/T 100 (M) 700 (M) ↓ not reported in F Bottero et al. 2005 F: atresia of previtellogenic oocytes & stroma Histo <2 2 yes alteration; M: inc. incidence testis alteration

The AFSS and other bioassays for antiandrogens 36

Table 1. cont.

Nominal Measured Species AA Exposure Duration Sex Endpoint(s) NOEC LOEC Effect Effect details Reference Comments Conc. (µg/L) Conc. (µg/L) MM CA Static 7d M/F 0.25, 1 no data plasma 11KT/T (M) < 0.25 0.25 ↓ plasma T/ E2 (F) < 0.26 1.25 ↓ VTG > 1 > 1 no Sharpe et al. 0.001, 0.010, MM CA Static 14d M/F * no data plasma 11KT/T (M) 0.01 0.1 ↓ 2004 0.100 * photoperiodically manipulated fish plasma T/ E2 (F) 0.001 (T) 0.01 (T) ↓ no effect in E2 levels VTG > 0.1 > 0.1 no

The AFSS and other bioassays for antiandrogens 37

Table 2. Synopsis of in vivo fish exposures to well known antiandrogens where females are stimulated by exogenous androgens.

"AA" Nominal "A" Nominal "AA" Species AA/A Exposure Duration (Measured) (Measured) Endpoint(s) NOEC a LOEC a "A" Effect b Effect details Reference Comments Effect a Conc.(μg/L) Conc.(μg/L) FHM FL/TB Flow 2w 400 (no data) 0.5 (no data) VTG mRNA > 400 > 400 no ↓ Ankley et al. 2004 SSC < 400 400 ↓ ↑ no quantitative data also genomic and proteomic FHM FL/TB Flow 48h 500 (560) 0.5 (0.47) VTG > 560 > 560 no no Garcia-Reyero et al. profiles analysed (same by 2009 Plasma E2 > 560 > 560 no ↓ Martyniuk et al. 2009) FHM CA/TB Flow 2w 20 (16), 200 (154) 0.5 (0.51)* SSC 16 154 ↓ ↑ Tubercle score *mean measured conc. from all treatments; also metabolite Ankley et al. 2010 profile analysed in urine by VTG > 154 > 154 no no Collette et al. 2010 FHM VZ/TB Flow 13d 200 (177), 700 (591) 0.5 (0.457)* SSC < 177 177 ↓ ↑ dec. tubercule score VTG > 591 > 591 no no high variability Martinović et al. *mean measured conc. from all Plasma E2/T > 591 > 591 no ↓ 2008 treatments mRNA > 591 > 591 no no 0.001 (0.0022), 0.01 FHM EE2/TB Flow 2w 0.5 (0.40)* SSC 0.0022 (F) 0.0089 (F) ↓ ↑ Tubercle score *mean measured conc. from all Ankley et al. 2010 (0.0089) treatments VTG > 0.0089 > 0.0089 no ↓ FHM BPA/TB Flow 2w 10 (12.6), 100 (94) 0.5 (0.52)* SSC 12.6 (F) 94 (F) ↓ ↑ Tubercle score *mean measured conc. from all Ankley et al. 2010 plasma VTG 12.6 94 ↑ ↓ treatments dec. nb segment (ray 3) & MF FL/ET Static 1-17dph 994 (no data) 1 (no data) SSC* < 994 994 (d9) ↓ ↑ retarded anal outgrowth Ogino et al. 2004 *gonopodium development dec. Shh expression & < 994 994 (d5) ↓ ↑ proliferating cells; qualitative assessment SB FL/DHT Flow 21d 500 (no data) 5 (no data) Spiggin < 500 500 ↓ ↑ Katsiadaki et al. SB FL/MT Static 21d 500 (no data) 1 (no data) Spiggin < 500 500 ↓ ↑ 2002b 0.1, 1, 10 (no SB FL/MT Static 21d 500 (no data) Spiggin < 500 500 ↓ ↑ data) 25 (17.3), 250 SB FL/MT Static 21d 0.5 (0.47)* Spiggin < 17.3 17.3 ↓ ↑ (263.7)* * measured conc. after addition SB VZ/MT Static 21d 25 (8.6), 250 (147.8)* 0.5 (0.47)* Spiggin < 8.6 8.6 ↓ ↑ of chemical; when MT was used Katsiadaki et al. at 5 µg/l the detection of the anti- 2006 SB LN/MT Static 21d 15 (8.7), 150 (127.3)* 0.5 (0.47)* Spiggin < 8.7 8.7 ↓ ↑ androgenic activity of all the compounds was not possible 15 (11.7), 150 SB FN/MT Static 21d 0.5 (0.47)* Spiggin < 11.7 11.7 ↓ ↑ (136.2)* SB FL/DHT Flow 21d 500 (no data) 5 (no data) Spiggin < 500 500 ↓ ↑ 1 (2.3), 10 (7.8), 50 Katsiadaki et al. **mean for all tanks, as similar SB FL/DHT Flow 21d (41.8), 125 (102.8), 5 (4.95)** Spiggin 2.3 7.8 ↓ ↑ 2006 (Lab 1, study measured conc. 250 (201.2), 500 1, OECD 2010) a as compared to the control group given a reference androgen agonist; b as compared to the control group The AFSS and other bioassays for antiandrogens 38

Table 2. cont.

"AA" Nominal "A" Nominal "AA" Species AA/A Exposure Duration (Measured) (Measured) Endpoint(s) NOEC a LOEC a "A" Effect b Effect details Reference Comments Effect a Conc.(μg/L) Conc.(μg/L) 5 (3.6), 25 (15.7), 100 Lab 4 study 2, OECD * mean for all tanks, as similar SB FL/DHT Flow 21d 5 (5.5)* Spiggin 15.7 31.4 ↓ ↑ (31.4), 250 (22) 2010 measured conc. 5 (2.1), 25 (10.2), 50 (22.3), 75 (32.3), 100 Lab 4 study 6, OECD * mean for all tanks, as similar SB FL/DHT Flow 21d 5 (3.5)* Spiggin 41.3 60.3 ↓ ↑ (41.3), 150 (60.3), 250 2010 measured conc. (96.3) 5 (4.1), 25 (14.3), 100 Lab 3 study 3, OECD * mean for all tanks, as similar SB FL/DHT Flow 21d 5 (5.4)* Spiggin 24.5 46.3 ↓ ↑ (24.5), 250 (46.3) 2010 measured conc. Lab 2 study 12, SB FL/DHT Flow 21d 250 (165) 5 (4.1) Spiggin < 165 165 ↓ ↑ OECD 2010 0.25 (0), 2.5 (0), 25 Lab 4 study 7, OECD * mean for all tanks, as similar SB VZ/DHT Flow 21d (2), 250 (15.1), 500 5 (4.6)* Spiggin 2 15.1 ↓ ↑ 2010 measured conc. (41.5) Lab 4 study 9, OECD * mean for all tanks, as similar SB VZ/DHT Flow 21d 10 (0.6), 100 (11.4) 5 (5.6)* Spiggin 0.6 11.4 ↓ ↑ 2010 measured conc. 0.25 (0), 2.5 (1), 25 Lab 4 study 8, OECD * mean for all tanks, as similar SB LN/DHT Flow 21d 5 (4.1)* Spiggin 1 7.6 ↓ ↑ (7.6), 250 (77) 2010 measured conc. 2 (0.7), 10 (4.8), 25 Lab 4 study 10, * mean for all tanks, as similar SB LN/DHT Flow 21d (12), 100 (58), 250 5 (6.8)* Spiggin 58 193.3 ↓ ↑ OECD 2010 measured conc. (193.3) Lab 4 study 11, * mean for all tanks, as similar SB LN/DHT Flow 21d 200 (73), 400 (94.7) 5 (4.5)* Spiggin < 73 73 ↓ ↑ OECD 2010 measured conc. 0.25 (0.04), 2.5 (0.3), Lab 3 study 4, OECD * mean for all tanks, as similar SB FN/DHT Flow 21d 5 (3.0)* Spiggin 0.3 3.6 ↓ ↑ 25 (3.6) 2010 measured conc.

0.1 (0.07), 0.25 (0.1), 0.625 (0.4), 1.56 (1.1), Lab 3 study 5, OECD * mean for all tanks, as similar SB FN/DHT Flow 21d 3.9 (2.4), 9.75 (6.5), 5 (4.6)* Spiggin 16.8 41.2 ↓ ↑ 2010 measured conc. 24.4 (16.8), 60 (41.2), 120 (88), 240 (177.6)

SB PRO/DHT Flow 21d 150 (no data) 5 (no data) Spiggin < 150 150 ↓ ↑

p, p' - SB Flow 21d 250 (no data) 5 (no data) Spiggin < 250 250 ↓ ↑ Katsiadaki et al. DDE/DHT unpublished data o,p' - SB Flow 21d 250 (no data) 5 (no data) Spiggin < 250 250 ↓ ↑ DDE/DHT a as compared to the control group given a reference androgen agonist; b as compared to the control group The AFSS and other bioassays for antiandrogens 39

Table 3. Comparison of the Hershberger and AFSS assays in term of responses to antiandrogenic chemicals .

"AA" Nominal "A" Nominal Fold ↓ at Assay AA/A a Endpoint(s) LOEC Reference Conc. range b Conc. b high conc. VP 0.3 5.1 SV 0.1 6.2 0.2 LABC 0.3 2.1 OECD 2006c COWS 0.3 3.4 FL/TP 0.1-10 GLANS 0.1 1.5 VP 0.1 7.3 SV 0.1 9.8 Hershberger Hershberger 0.4 LABC 0.1 2.7 OECD 2006c COWS 0.1 4.3 GLANS 0.3 1.6 FL/DHT 500 5 Spiggin 500 5600 Katsiadaki et al. 2002b

AFSS 1-500 Spiggin 50 726.6 OECD 2010 VP 3 1.9 SV 10 5 0.2 LABC 10 2 OECD 2008 COWS 10 2.5 VZ/TP 3-100 GLANS 10 1.4 VP 30 2.7 SV 30 2.9 Hershberger 0.4 LABC 30 1.9 OECD 2008 COWS 30 1.8 GLANS 100 1.3

VZ/DHT 0.25-500 5 Spiggin 100 820.3 OECD 2010 AFSS VP 100 1.7 SV 30 1.9 3-100 0.4 LABC 30 1.5 OECD 2008 COWS 100 1.3 GLANS 100 1.1 VP 10 2.4 SV 10 2.8 LN/TP 0.2 LABC 100 1.8 COWS 100 1.6 Hershberger 10 and 100 GLANS 100 1.2 OECD 2007 VP 10 1.9 SV 10 2.2 0.4 LABC 10 1.7 COWS 100 1.7 GLANS 100 1.1

LN/DHT 0.25-400 5 Spiggin 250 10 OECD 2010 AFSS VP 15 2.5 FN/TP 15 and 30 0.2 SV 15 2 Tamura et al. 2001 LABC 15 1.5 Hershberger

FN/DHT 0.1-240 5 Spiggin 60 460.7 OECD 2010 AFSS a AA/A: antiandrogen and androgen b in mg/kg/day for the Hershberger assay and in µg/L for the AFSS assay

The AFSS and other bioassays for antiandrogens 40

Table 3. cont.

"AA" Nominal "A" Nominal ↓ Fold at Assay AA/A a Endpoint(s) LOEC Reference Conc. range b Conc. b high conc. VP 30 2.1 SV 30 2.8 3-100 0.2 LABC 30 1.6 COWS 30 1.8 GLANS 100 1.2 OECD 2008 VP 50 2.7 SV 50 2.9 5-160 0.4 LABC 16 2 COWS 16 2.2 p, p' -DDE/TP GLANS 50 1.4 VP 16 1.1 SV 16 4.4 0.2 LABC 160 2.3 COWS 160 2.5 16 and 160 GLANS 160 1.6 OECD 2007 VP 16 2.3 SV 160 3.3 0.4 LABC 16 2.1 COWS 16 2.2 GLANS 16 1.3 p, p' -DDE/TP 250 5 Spiggin 250 46 Katsiadaki et al. unpublished o, p' -DDE/TP 250 5 Spiggin 250 69.1 data 3-100 0.4 VP 10 2.8 SV 10 2.6 PRO/TP LABC 10 1.8 OECD 2008 COWS 10 2.1

Hershberger Hershberger GLANS 30 1.2 Katsiadaki et al. unpublished PRO/TP 150 5 Spiggin 150 56.9

AFSS data VP > 160 1.1 SV > 160 1.1 0.2 LABC > 160 1 COWS > 160 1 NP 160 GLANS > 160 1 OECD 2007 VP > 160 1.1 SV > 160 0.97 Hershberger Hershberger AFSS 0.4 LABC > 160 1 COWS > 160 1 GLANS > 160 1 Katsiadaki et al., unpublished NP/DHT 25-500 5 Spiggin 500 3 AFSS data a AA/A: antiandrogen and androgen b in mg/kg/day for the Hershberger assay and in µg/L for the AFSS assay

The AFSS and other bioassays for antiandrogens 41

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