Aquatic Toxicology 144–145 (2013) 19–25

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Estrogenic effect of the phytoestrogen biochanin A in zebrafish, Danio rerio, and brown trout, Salmo trutta Henrik Holbech, Kristoffer D. Schröder, Marie L. Nielsen, Nanna Brande-Lavridsen, Bente Frost Holbech, Poul Bjerregaard ∗ Institute of Biology, University of Southern Denmark, Campusvej 55, DK-5230 Odense, Denmark

a r t i c l e

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Article history: Received 10 July 2013 Received in revised form 1 September 2013 Accepted 5 September 2013 Keywords: Biochanin A Vitellogenin Sex ratio Brown trout Zebrafish

a b s t r a c t Isoflavones with estrogenic activity produced in Fabaceae plants are known to leach from agricultural areas to freshwater systems, but the effect of waterborne isoflavones in fish has not been thoroughly characterized. Therefore, the estrogenic effect of waterborne biochanin A was investigated in zebrafish (Danio rerio) and juvenile brown trout (Salmo trutta). Exposure of juvenile brown trout to 10 ␮g biochanin A L−1 or higher caused marked vitellogenin induction after 9–10 days of exposure and so did exposure to 186 ␮g biochanin A L−1 for 6 h. Following 8 d of exposure, a NOEC for induction of vitellogenin production in male zebrafish was 70 and LOEC 114 ␮g biochanin A L−1 . Exposure to 209 ␮g biochanin A L−1 from hatch to 60 days post hatch (dph) caused a skewing of the sex ratio toward more phenotypic female zebrafish, but did not cause induction of vitellogenin in male and undifferentiated fish. In conclusion: (1) biochanin A elicits estrogenic effects in trout at environmentally realistic concentrations, (2) brown trout plasma vitellogenin concentrations respond to lower biochanin A exposure concentrations than vitellogenin concentrations in zebrafish homogenates and (3) concerning vitellogenin induction, the hypothesis should be tested if short term tests with zebrafish may show a higher sensitivity than partial life cycle tests. © 2013 Elsevier B.V. All rights reserved.

1. Introduction Feminization of male fish due to discharge of estrogenic substances to the aquatic environment has been known for a couple of decades and in most cases these effects have been attributed to the presence of natural [17␤-estradiol (E2) and estrone (E1)] and synthetic [17␣-ethinylestradiol (EE2)] estrogens discharged from waste water treatment plants (WWTP) (Sumpter, 2005). Feminization of male fish has, however, also been observed in streams not affected by discharges from WWTPs, and discharges of estrogenic activity from other sources such as agriculture (Kjaer et al., 2007; Matthiessen et al., 2006) and houses in the open land (Stuer-Lauridsen et al., 2006) have been investigated. In their investigation of estrogenic discharges from dairy farms, Matthiessen et al. (2006) could not account for all of the estrogenic activity measured by the presence of natural estrogens alone and they suggested that some of the estrogenic activity might possibly be attributed to releases of phytoestrogens, potentially from silage. Phytoestrogens are intermediary metabolites produced especially in Fabaceae plants such as clover, alfalfa, peas (mainly

∗ Corresponding author. Tel.: +45 65502456; fax: +45 65502786. E-mail address: [email protected] (P. Bjerregaard). 0166-445X/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.aquatox.2013.09.006

biochanin A and formononetin) and soy (mainly daidzein and genistein) (reviewed by Price and Fenwick, 1985). These compounds function in microbial recruitment as a signal for nitrogenfixing bacteria to form nodules and possibly also help defend the plants against herbivores and pathogens (Fox et al., 2004; Wynne-Edwards, 2001). The estrogenic activity of phytoestrogens (resulting in infertility) in vertebrates was first recognized in 1946 as a syndrome known as clover disease in sheep (Bennetts et al., 1946). Phytoestrogens reach the aquatic environment after release from the cultivated plant to the soil water and runoff from manuretreated soils (Burnison et al., 2003; Erbs et al., 2007; Hoerger et al., 2009). Phytoestrogens have also been detected in waste streams from plant-processing industries including biofuel manufacturers, paper and pulp mills (Lundgren and Novak, 2009), as well as domestic effluents (Bacaloni et al., 2005; Farre et al., 2008). Phytoestrogens have been detected in freshwater ecosystems around the world, typically in the ng per liter range (e.g. Hoerger et al., 2009; Lundgren and Novak, 2009; Ribeiro et al., 2009), but in some cases up to the ␮g per liter range (Kawanishi et al., 2004; Lundgren and Novak, 2009). Phytoestrogens, such as biochanin A and its metabolite genistein are able to elicit estrogenic effects in fish and affect their reproductive functions. The major part of this knowledge relates to the

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use of soy containing diets in aquaculture (e.g. Latonnelle et al., 2002) and relatively little is known about the estrogenic potency of waterborne biochanin A. The purpose of the present investigation was to derive concentration–response relationships for the estrogenic effect of biochanin A in zebrafish and brown trout using the estrogendependent induction of vitellogenin and change in phenotypic sex in juvenile zebrafish as biomarkers. 2. Materials and methods 2.1. Chemicals Biochanin A (CAS# 491-80-5), DMSO (CAS# 67-68-5) and 17␤estradiol (CAS# 50-28-2) were from Sigma–Aldrich, Vallensbæk Strand, Denmark. 2.2. Experimental animals Adult zebrafish were bought from a local supplier and adapted to laboratory conditions for several weeks during which they were fed with TetraMin (Tetra GmbH, Melle, Germany). Sexually immature brown trout were obtained from two commercial fish farms in Denmark (Vork & Hvilested). The trout were acclimated for 3 days in the laboratory before experiments began. During this period they were kept in 200 L tanks with a supply of recirculating groundwater and fed with commercial trout pellets (Aller Aqua, Denmark). 2.3. Exposure systems The exposure systems were flow through test systems with 8 L glass aquaria (containing 6 L of water) for the zebrafish and 112.5 L stainless steel tanks for the brown trout. Water exchange was 18 L per 24 h for the zebrafish and between 130 and 173 L per 24 h for the trout. Micra or Secci circulation pumps positioned in the trout tanks were used to keep oxygen level above 80% saturation and to ensure complete and uniform dispersal of the test compounds. The trout were supplied with tap water (unchlorinated groundwater), the adult male zebrafish with 1 part tap water and 3 parts deionized water and the developing zebrafish with a 50:50 mixture of tap water and deionized water. Administration of water and test compound was controlled by two peristaltic pumps (Ole Dich Instrumentmakers, Denmark). Stock solutions of the test compounds were prepared in 50% DMSO before administration. Dilution water and stock solution were mixed to the desired concentration before entering the aquaria and the maximum DMSO concentration in the exposure water was 50 ␮L L−1 . Photoperiods were 12:12 for both species. Water temperature was maintained at 27 ± 2 ◦ C for the zebrafish and 14 ± 1 ◦ C for the trout. In both the initial brown trout and adult male zebrafish experiment, an estrogen reference group exposed to nominal concentrations of 50 (trout) and 200 (zebrafish) ng E2 L−1 was employed to ensure that the fish were responsive to estrogens. In the sexual development test, four replicate aquaria of control, solvent control and biochanin A exposures were used; in the remaining experiments one tank was used per exposure concentration. 2.4. Sexual development in zebrafish Late in the afternoon spawning chambers with artificial plants were placed in the aquarium containing approximately 50 adult zebrafish and removed again the following morning. The newly fertilized eggs were collected and randomly divided into groups

of 100 and placed in 400 mL glass vessels. One day post fertilization (dpf) unfertilized eggs were removed and replaced by fertilized eggs from a reservoir and a total of 90 fertilized eggs were transferred to each of 20 exposure tanks. At three days post hatch (dph) the larvae were fed three times daily with Sera micron powdered food for fry (Heisenberg, Germany) and at 23 dph it was supplemented with TetraMin® Baby (Tetra GmbH, Melle, Germany). At 26 dph dry food was solely comprised of TetraMin® Baby. The baby food was substituted by crushed TetraMin® flakes (Tetra GmbH, Melle, Germany) at 45 dph. In addition newly hatched Artemia sp. nauplii (Inter Yyba GmbH, Germany) were supplied once or twice daily from 3 dph. The exposure (19.3 ± 1.1, 61 ± 3.8 and 209 ± 14 ␮g biochanin A L−1 ) took place from one dpf until 60 dph. Aquaria were aerated from 7 dph. During exposure, conductivity, oxygen saturation and temperature were measured twice per month. Conductivity ranged from 220 to 250 ␮S and average oxygen saturation ranged from 75% to 78%. At 60 dph all fish were euthanized in an overdose of buffered MS222. Fish were randomly selected evenly from the four replicates for histological analyses and determination of vitellogenin. A few of the samples were lost during fixation; the actual number of fish analyzed is given in Fig. 1. 2.5. Short term exposure of adult male zebrafish In two experiments, groups of 9–12 adult, male zebrafish were exposed to various (nominal: 63, 125, 250, 500 and 8, 16, 32, 64, 128, 256 ␮g L−1 ) concentrations of biochanin A for 8 days and vitellogenin induction was used as estrogenic biomarker. The fish were not fed during the exposure. 2.6. Brown trout experiments In two experiments, groups of 13–17 juvenile brown trout were exposed to various concentrations (nominal: 38, 75, 150, 320 and 4, 8, 16, 32, 64 ␮g L−1 ) of biochanin A for 9 or 10 days and plasma vitellogenin was determined as estrogenic endpoint. In a third experiment, juvenile brown trout were exposed to nominal concentrations of 100, 320 and 1000 ␮g L−1 of biochanin A in 6 h pulses

Fe ma le s

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Biochanin A concentration (μ μg L ) Fig. 1. Danio rerio. Percentage females, males and undifferentiated zebrafish after exposure to biochanin A in the water from 0 to 60 days post hatch. n given in histogram inserts. * indicates P < 0.05. Actual concentrations (mean ± SEM; n = 80) of biochanin A are shown.

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and vitellogenin concentrations in the plasma were determined 3 days after the pulse exposure. At the end of each experiment, the fish were anesthetized in MS-222 and length and weight (average in the 3 experiments: 17.2 ± 0.5, 2.7 ± 0.1 and 4.3 ± 0.2 g) were recorded. Blood samples were taken and kept in heparinised tubes. The blood was centrifuged for 10 min at 13,400 rpm and plasma was stored at −80 ◦ C until the vitellogenin concentration was measured.

2.7. Quantification of biochanin A Water samples were passed through 0.45 ␮m poylvinylidene difluoride filters (Frisenette, Denmark) before actual exposure concentrations were determined by high-performance liquid chromatography–tandem mass spectrometry (A 1200 Series HPLC and a 6410 Triple Quad LC/MS, both Agilent Technologies). Six ␮L water sample were injected in the HPLC-MS/MS and conditions were as follows: column Zorbax Eclipse XDB C18 4.6 × 50 mm, 1.8 ␮m Rapid Resolution HT, column temperature 70 ◦ C, A: 1% formic acid and 20% acetonitril and B: Acetonitril with a gradient from 40% to 70% over 2 min, held at 70% for 1 min followed by equilibration at 40%, flow 1.0 mL/min, Electrospray Ionization in positive ion mode, drying gas flow 11.0 L/min, nebulizer pressure 35 psig, drying gas temperature 350 ◦ C and capillary voltage of 4000 V. Fragmentor voltage was 110 V and collision energies were 45 V (quantifier) and 35 V (qualifier). The mass-to-charge ratios (m/z) of precursor, quantifier and qualifier ions were 285.2, 213.0 and 152, respectively. The standards were prepared by dissolving biochanin A (Sigma D2016) in 20% MeOH. The detection limit was approximately 0.4 ␮g biochanin A L−1 .

3. Results 3.1. Biochanin A water concentrations The actual concentrations of biochanin A are presented as numbers on the x-axis in Figs. 1–5. Actual concentrations were generally lower than nominal concentrations; in the first zebrafish experiment, a pump failure gave an actual concentration of 3.8 ␮g L−1 instead of the nominal 250 ␮g L−1 . 3.2. Sexual development in zebrafish In the control groups, 31.7% survived from eggs to 60 d and the average weight was 66.8 mg; there were no statistically significant differences in the exposure groups. Approximately 1⁄3 of the fish did not exceed a weight of 35 mg after 60 days irrespective of exposure (Fig. 1) and these fish were sexually undifferentiated. The control and solvent control groups and the group exposed to 19 ± 1 ␮g biochanin A L−1 contained more than twice as many males than females (Fig. 1). The group exposed to 209 ± 14 ␮g biochanin A L−1 had a significantly increased number of females (Fig. 1), whereas only a slight trend was seen in the group exposed to 61 ± 4 ␮g biochanin A L−1 . Vitellogenin concentrations were elevated in the male and small fish relative to the control in the groups exposed to 19 ± 1 ␮g biochanin A L−1 for 60 days (Fig. 2), whereas there was no difference between the vitellogenin levels in the control groups

2.9. Data handling and statistical treatment Control and solvent control were compared: if significantly different, the solvent control was used; if not, the controls were combined. Vitellogenin data: Normality and variance homogeneity were checked and data were logarithmically transformed if necessary. If normality and variance homogeneity could be obtained, a one-way ANOVA followed by a parametric Bonferroni–Holm test was used. If not, a non-parametric Kruskal–Wallis ANOVA followed by Dunn’s test was used. Sex ratio data: Differences between groups were tested using the Chi2 -test. Significance level: ˛ = 0.05.

-1

Vitellogenin level (ng g wet weight)

2.8. Vitellogenin analysis and histology Vitellogenin levels were determined by species-specific Enzyme Linked Immuno Sorbent Assays (ELISA) developed for zebrafish (Holbech et al., 2001) and brown trout (Bjerregaard et al., 2006). Zebrafish were cut into three sections, head, trunk (containing the gonads) and tail; vitellogenin levels were determined in the homogenates of the tail and head sections as described by Holbech et al. (2001, 2006) and Morthorst et al. (2010). In brown trout, vitellogenin levels were determined in plasma according to Bjerregaard et al. (2006). The detection limit was 0.2 ng vitellogenin mL−1 for both brown trout plasma and zebrafish homogenate. Intra- and interassay coefficients of variation have been determined to 8.1% and 16.7% for the trout ELISA (Bjerregaard et al., 2006) and 5.8% and 10.4% for the zebrafish ELISA (Holbech et al., 2001). The gonads were dissected out from the trunk section, and the sex of the zebrafish in the sexual development test was determined by histological examination of gonads as described by Kinnberg et al. (2007).

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Biochanin A concentration (μg L ) Fig. 2. Danio rerio. Vitellogenin concentrations in zebrafish exposed to biochanin A in the water from 0 to 60 days post hatch. Fifty and 90 percentile box plot with median (♦) and outliers (䊉). The number of fish in each group is shown above the x-axis. ** indicates P < 0.01. Actual concentrations (mean ± SEM; n = 80) of biochanin A are shown.

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4. Discussion Phytoestrogens may have both estrogenic and anti-estrogenic effects depending on the specific tissue and the concentration of circulating endogenous estrogens (Benassayag et al., 2002), but in the experiments reported here, the effects of biochanin A were

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Exposure of juvenile brown trout to concentrations above approximately 10 ␮g biochanin A L−1 resulted in markedly elevated concentrations of vitellogenin after 9 or 10 days exposure (Fig. 4). Vitellogenin levels in the groups exposed to 1.2 ± 0.2 and 3.8 ± 1.0 ␮g biochanin A L−1 were significantly higher than the levels in the control groups, but the induction was limited and vitellogenin levels in the group exposed to 1.8 ± 0.5 ␮g biochanin A L−1 did not deviate from that of the controls. Exposure to 6 h pulses of 186 ± 13 and 651 ± 12 ␮g biochanin A L−1 resulted in elevated vitellogenin levels, whereas exposure to 48 ± 5 ␮g biochanin A L−1 did not (Fig. 5).

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Fig. 4. Salmo trutta. Plasma vitellogenin concentrations in juvenile brown trout exposed to biochanin A in the water for 9 (A) or 10 (B) days. Symbols as explained in Fig. 3. E2 in A: exposed to 50 ng E2 L−1 . Actual concentrations (mean ± SEM; n = 15 in A and 13 in B) of biochanin A are shown. *, ** and *** indicate P < 0.05, 0.01 and 0.001, respectively.

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3.4. Juvenile brown trout tests

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Vitelloge enin level (ng mL plasma a)

Exposure of adult, male zebrafish to concentrations above approximately 100 ␮g biochanin A L−1 resulted in elevated concentrations of vitellogenin (Fig. 3).

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Biochanin A concentration (μg L )

and the groups exposed to 61 ± 4 and 209 ± 14 ␮g biochanin A L−1 (Fig. 2). 3.3. Vitellogenin induction in adult zebrafish

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Biochanin A concentration (μg L ) Fig. 3. Danio rerio. Vitellogenin concentrations in male zebrafish exposed to biochanin A in the water for 8 days. The number of fish in each group is shown above the x-axis. Fifty and 90 percentile box plot with median (♦) and outliers (䊉). Actual concentrations of biochanin A are shown (mean ± SEM; n = 4 in A and 5 in B). E2 in A: Exposed to 200 ng E2 L−1 . *, ** and *** indicate P < 0.05, 0.01 and 0.001, respectively.

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Biochanin A concentration (μg L ) Fig. 5. Salmo trutta. Plasma vitellogenin concentrations in juvenile brown trout 3 days after exposure to 6 h pulses of biochanin A. Symbols as explained in Fig. 3. Actual concentrations (mean ± SEM; n = 3) of biochanin A are shown. *** indicates P < 0.001.

H. Holbech et al. / Aquatic Toxicology 144–145 (2013) 19–25

consistent with its role as an estrogen receptor (ER) agonist inducing vitellogenin and skewing phenotypic sex toward females; the investigation was not designed to reveal potential anti-estrogenic effects. 4.1. Vitellogenin induction. Exposure to biochanin A concentrations around and above 10 ␮g L−1 causes markedly elevated vitellogenin production in the juvenile brown trout and concentrations between 1 and 10 ␮g L−1 may cause inductions that may or may not be statistically significant. The ability of biochanin A to induce vitellogenin synthesis in fish has been shown before (Katchamart et al., 2002), but in contrast to the present study, the effect was obtained after injection of fairly high doses (25 with 50 mg kg−1 body weight in juvenile rainbow trout) of biochanin A. Since phytoestrogens have been detected in freshwater ecosystems around the world, in some cases up to the ␮g per liter range (Kawanishi et al., 2004), the results of the present investigation show that phytoestrogens should always be considered if estrogenic effects are detected which cannot be explained by the presence of steroid hormones (Matthiessen et al., 2006). Chemicals may often be present in streams as short pulses associated with precipitation incidents (e.g. Kjaer et al., 2007). It takes 10–20 times higher concentrations of biochanin A to elicit a vitellogenin response in brown trout after 6 h than after 9–10 d exposure. This is comparable to the effects of E2 where the EC50 after 10 d exposure is 15 ng L−1 (Bjerregaard et al., 2008), whereas it takes E2 concentrations at or above 150–200 ng L−1 to elicit a response after 6 h exposure (Knudsen et al., 2011). The concentrations of biochanin A required to induce a response after the 6 h pulse exposure are higher than will probably ever be recorded in the environment. In zebrafish, concentrations of biochanin A at or above 100 ␮g L−1 are required to elicit a vitellogenin response in adult males after 8 d exposure. This renders the brown trout more than 10 times as sensitive as the zebrafish when the estrogenic effects of biochanin A on vitellogenin induction are considered, keeping in mind that vitellogenin levels are determined in plasma and body homogenate of the trout and zebrafish, respectively. However, if comparison of species sensitivity toward any given chemical has to be made, real life data to do so will very often have been acquired in the form of plasma vitellogenin concentrations from wild species such as trout or flounder and body homogenates from zebrafish (e.g. from OECD TG 234). Zebrafish are often used in the laboratory to generate data for hazard and risk assessment of chemicals; in chemical regulation, it is useful to have in mind that other species may be more sensitive to exposure to chemicals than zebrafish. At exposure to natural (E2) or synthetic steroids (EE2) there is, however, little difference between the sensitivity of brown trout (Bjerregaard et al., 2008) and zebrafish (Rose et al., 2002) as far as vitellogenin induction is concerned. The mechanism underlying the difference in the response patterns for the steroids and the phytoestrogen is not known, but Van den Belt et al. (2003) reviewed differences between species in the effects of phytoestrogens, suggesting that vitellogenin production is dependent on species but also varies with the different life stages of the individuals tested. In our short term experiments juvenile brown trout and adult, male zebrafish were used and reproductive status and thereby natural estrogen differed, which can account for some of the difference in the response to biochanin A. Exposure to 4-tert-octylphenol and 4-nonylphenol induced vitellogenin synthesis in juvenile rainbow trout but not or little in adult zebrafish under similar exposure conditions while induction due to several other estrogenic chemicals showed no species differences (Van den Belt et al., 2003). Also, trout and sturgeon exhibited a 50-fold difference in sensitivity to genistein with regard to vitellogenin synthesis (Gontier-Latonnelle

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et al., 2007; Latonnelle et al., 2002). This was not due to differences in ER affinity or to the sensitivity of the vitellogenin induction pathway (Gontier-Latonnelle et al., 2007; Latonnelle et al., 2002); plasma levels of genistein differed after exposure and differences in metabolism and excretion of the compound were suggested. It is unexpected that the male and undifferentiated zebrafish did not show elevated vitellogenin levels after 60 d exposure to 209 ␮g biochanin A L−1 in the test covering sexual development, when NOEC- and LOEC-values for vitellogenin induction in the adult male zebrafish after 8 d exposure were 70 and 114 ␮g biochanin A L−1 , respectively. Small undifferentiated fish (34 dph) have previously been shown to respond to xenoestrogenic exposure (Holbech et al., 2006) so we would have expected an induction of vitellogenin in the present study. It has, however, been shown that juvenile fish are less sensitive to E2 than adult males with regard to vitellogenin induction (Brion et al., 2004), but it is still unexpected that exposure concentrations capable of altering the sex ratio are not necessarily causing induction of vitellogenin in a 60 d test. Adaptation of some kind to the estrogenic exposure may be an explanation, but this phenomenon certainly needs further investigation. It is a generally accepted principle that endpoints in partial and full life cycle tests are more sensitive to the effects of chemicals than endpoints in short term tests, but the results of the present investigation indicate that we need to know more about the potential role of adaptation in long-term tests. The lack of differences in vitellogenin concentrations in the female 60 dph zebrafish is not as surprising as that in the males, since biochanin A is not a very potent estrogenic substance compared with natural estrogen activity in adult females. The overall survival (∼32%) in the 60 d study was not related to exposure but somewhat lower than recommended in the recently developed test guideline (OECD, 2011a) (80% hatch- and 70% posthatch survival – in total 56%) and the average weight of the fish (∼67 mg) only approached the recommended value of 75 mg. The high mortality and low growth were probably caused by a relatively high number of fertilized eggs (90 per replicate) in the test system. 4.2. Sex ratio We observed a greater proportion of phenotypically female zebrafish individuals as a result of aqueous exposure to 209 ␮g biochanin A L−1 from 0 to 60 dph. Sexual differentiation in zebrafish is very sensitive to exposure to exogenous hormones and zebrafish populations can be made all female by exposure to natural and synthetic steroid estrogens (Holbech et al., 2006; Nash et al., 2004) during the period of sexual differentiation and all male or male biased populations can be produced by exposure to androgens (Morthorst et al., 2010; Orn et al., 2003) or aromatase inhibitors (Holbech et al., 2012; Kinnberg et al., 2007; Thorpe et al., 2011). Chemicals (e.g. 4-nonylphenol, 4-tert octylphenol and 4-tert pentylphenol) with weaker estrogenic effects than the steroid hormones have previously been shown to have the ability to alter the sex ratio in zebrafish (OECD, 2011b,c). Only 60% of the fish differentiated sexually in the experiment and the percentage was not related to the exposure. The sex ratios in the control groups were skewed toward a greater proportion of males and contained less than 20% females. Such skewed control sex ratios have been noticed in other studies (Hensley and Leung, 2010; Larsen et al., 2008; Orn et al., 2006), and in a recent experiment (Brown et al., 2012) a significant male-biased (72%) sex ratio was observed in the most inbred line of zebrafish lines studied. In the present study, the parental fish were purchased from a local supplier who does not register the breeding history of the fish so unfortunately we cannot connect the inbreeding status to the sex ratio. Based on the results from other studies (Brown et al., 2012; Thorpe et al., 2011) there is no reason to assume that the skewed sex

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ratio is caused by slower development among the females such that the undifferentiated fish would have turned into an overweight of females if they were allowed to differentiate or if female mortality was over-represented. A theory about size and sexual development has been proposed by Lawrence et al. (2008) and supported by Gomez-Requeni et al. (2010). These authors propose that as larger females potentially produce more gametes than smaller females, individuals that grow slowly are more likely to have higher reproductive success as small males than as small females. Therefore, a slow growth would lead to a male biased sex ratio as observed in the present experiment. 4.3. Criteria for endocrine disrupting chemicals For a chemical to be designated ‘endocrine disrupting’ there is general agreement that both endocrine specific and ‘adverse’ endpoints must be affected. Both vitellogenin induction and sex ratio changes are considered endocrine specific, whereas only the sex ratio change is also considered ‘adverse’ since vitellogenin induction on its own does not necessarily have an impact on population parameters. Vitellogenin induction would normally be considered an equally or more sensitive endpoint than a change in sex ratio, but in the present investigation an exposure range that caused vitellogenin induction in a short term experiment with adult males and skewed sex ratio in juvenile zebrafish in a partial life cycle test, caused no vitellogenin induction in the latter. This conflict shows that we need to know more about the role of potential adaptation to estrogenic exposure in long term experiments. Authors’ contributions HH and PB planned and supervised the series of experiments. KDS carried out the zebrafish experiments. MLN carried out the brown trout experiments. BFH developed and performed the chemical analyses. NB-L performed part of the histological examination of the zebrafish and participated in the drafting of the manuscript. All of the authors approved the manuscript. Acknowledgment This investigation was supported by grants from the Danish Natural Science Research Council. References Bacaloni, A., Cavaliere, C., Faberi, A., Foglia, P., Samperi, R., Lagana, A., 2005. Determination of isoflavones and coumestrol in river water and domestic wastewater sewage treatment plants. Analytica Chimica Acta 531, 229–237. Benassayag, C., Perrot-Applanat, M., Ferre, F., 2002. Phytoestrogens as modulators of steroid action in target cells. Journal of Chromatography B – Analytical Technologies in the Biomedical and Life Sciences 777, 233–248. Bennetts, H., Underwood, E., Shier, F., 1946. A specific breeding problem of sheep on subterranean clover pastures in Western Australia. Australian Veterinary Journal 22, 2–12. Bjerregaard, L.B., Madsen, A.H., Korsgaard, B., Bjerregaard, P., 2006. Gonad histology and vitellogenin concentrations in brown trout (Salmo trutta) from Danish streams impacted by sewage effluent. Ecotoxicology 15, 315–327. Bjerregaard, P., Hansen, P., Larsen, K.J., Erratico, C., Korsgaard, B., Holbech, H., 2008. Vitellogenin as a biomarker for oestrogenic effects in brown trout, Salmo trutta: laboratory and field investigations. Environmental Toxicology and Chemistry 27, 2387–2396. Brion, F., Tyler, C.R., Palazzi, X., Laillet, B., Porcher, J.M., Garric, J., Flammarion, P., 2004. Impacts of 17 beta-estradiol, including environmentally relevant concentrations, on reproduction after exposure during embryo-larval-, juvenile- and adult-life stages in zebrafish (Danio rerio). Aquatic Toxicology 68, 193–217. Brown, A.R., Bickley, L.K., Ryan, T.A., Paull, G.C., Hamilton, P.B., Owen, S.F., Sharpe, A.D., Tyler, C.R., 2012. Differences in sexual development in inbred and outbred zebrafish (Danio rerio) and implications for chemical testing. Aquatic Toxicology 112–113, 27–38. Burnison, B.K., Hartmann, A., Lister, A., Servos, M.R., Ternes, T., Van der Kraak, G., 2003. A toxicity identification evaluation approach to studying estrogenic

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