Trenbolone causes irreversible masculinization of zebrafish at environmentally relevant concentrations

Aquatic Toxicology 98 (2010) 336–343

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Trenbolone causes irreversible masculinization of zebrafish at environmentally relevant concentrations Jane E. Morthorst ∗ , Henrik Holbech, Poul Bjerregaard Institute of Biology, University of Southern Denmark, Campusvej 55, DK-5230 Odense M, Denmark

a r t i c l e

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Article history: Received 23 January 2010 Received in revised form 5 March 2010 Accepted 8 March 2010 Keywords: Zebrafish Sex ratio Masculinization Recovery Trenbolone Androgens

a b s t r a c t Feminization of fish caused by certain estrogenic compounds e.g. 17␣-ethinylestradiol (EE2) has been shown to be partly reversible. So far it has not been studied if this applies for androgenic compounds too. The androgenic steroid trenbolone acetate (TbA) is used as growth promoter in beef cattle in the United States, South America, and Australia. TbA metabolites are stable in animal waste and have been detected in surface waters associated with feedlot areas and studies on both fish and mammals have demonstrated a strong androgenic effect of those metabolites. Zebrafish (Danio rerio) were exposed to environmentally relevant concentrations of the TbA metabolite 17␤-trenbolone from 0 to 60 days post-hatch (dph) and either sacrificed at 60 dph, transferred to clean water for 170 days or kept in exposure for 170 days. At 60 dph gonadal histology and vitellogenin analyses revealed all-male populations in groups exposed to 15.5 and 26.2 ng/L, and at 9.2 ng/L a skewed sex ratio towards males was observed. After the depuration period no sign of reversibility was observed. Environmentally relevant concentrations of 17␤-trenbolone cause a strong and irreversible masculinization of zebrafish and that raises concern about the effects of androgenic discharges in the aquatic environment. In addition this study also aids in understanding of the so far unknown sex determination process in zebrafish. © 2010 Elsevier B.V. All rights reserved.

1. Introduction The synthetic androgen trenbolone acetate (TbA) is used as a growth promoter in beef cattle. Within the animals TbA is metabolized into 17␣-trenbolone and 17␤-trenbolone and both metabolites are excreted with feces and urine. The metabolites are fairly stable in animal waste and the environment (Pottier et al., 1981; Schiffer et al., 2001) and have been found in surface waters associated with beef feedlots (Durhan et al., 2006). The usage of fish as a vertebrate model to reveal endocrine disrupters is justified by the striking similarity of the hormonal system among vertebrates (Ankley and Johnson, 2004; Mills and Chichester, 2005) and compared to other vertebrates, fish are easier and cheaper assigned to genetic and embryological manipulation. Zebrafish (Danio rerio) are widely used as they are very robust and spawn continuously (Spence et al., 2008). Lately the entire zebrafish genome has been sequenced but still it is unclear how sexual development is triggered and regulated as neither sex chromosomes nor sex-determining genes have been found in zebrafish (Traut and Winking, 2001; Wallace and Wallace, 2003; Jorgensen et al., 2008).

∗ Corresponding author. Tel.: +45 6550 2760; fax: +45 6550 2786. E-mail address: [email protected] (J.E. Morthorst). 0166-445X/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.aquatox.2010.03.008

The feminizing effects of moderate concentrations of the synthetic estrogen 17␣-ethinylestradiol (EE2) on zebrafish have been demonstrated to be partly reversible (Maack and Segner, 2004; Nash et al., 2004; Schafers et al., 2007) but reversibility of effects caused by androgens has not been studied. The primordial germ cells (PGCs) have been shown to play a role in zebrafish sexual development. Zebrafish, in which PGCs were inactivated by knockdown, all developed as sterile males (Slanchev et al., 2005). New studies have confirmed development of an all-male population by ablation of PGCs in zebrafish embryos (Saito et al., 2008; Siegfried and Nusslein-Volhard, 2008). In medaka (Oryzias latipes) those results were supported as the essentiality of germ cells in sexual dimorphism of medaka gonads was demonstrated (Kurokawa et al., 2007). Seen in the light of the above mentioned experiments it would be important for the overall understanding of sex determination in zebrafish to know if androgenic effects are also reversible. With the results from Slanchev et al. (2005) and Kurokawa et al. (2007) in mind we hypothesized that masculinizing effects of trenbolone would be irreversible. The objective of this study was to investigate the reversibility of the masculinizing effects of environmentally relevant concentrations of 17␤-trenbolone, which is a TbA metabolite, in zebrafish. The obtained results would provide information about the environmental influences of androgenic compounds discharged into surface water and the unknown features of sex determination in

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zebrafish. Gonad histology and vitellogenin analyses were used as endocrine endpoints. 2. Materials and methods 2.1. Chemicals 17␤-Trenbolone (4,9,11-estratrien-17␤-ol-3-one) was purchased from Steraloids Inc. (Newport, USA). Acetone (96%) was purchased from Sigma–Aldrich (Vallensbæk Strand, Denmark) and diluted to 50% with ASTM type 1a water. Stock solutions of 17␤trenbolone dissolved in 50% acetone were prepared every third week. 2.2. Animals Parent fish were bought from a local supplier and adapted to laboratory conditions for several weeks. Spawning chambers with artificial plants were used to collect the eggs. After spawning had ended the newly fertilized eggs were collected and counted, divided into groups of 100 and put into 400 mL glass dishes (27 ± 1 ◦ C). One day post-fertilization (dpf) unfertilized eggs were removed and replaced by fertilized eggs from a reservoir, hence in each dish the total number of viable eggs was still 100. A total of 100 eggs were transferred to each exposure tank. The experiment was in general performed according to the OECD TG 210 ‘Fish Early Life Stage Toxicity Test’. 2.3. Exposure A flow through test system with 8 L aquaria containing 6 L of water and a water exchange of 18 L/day was used during exposure and depuration periods. Larvae were kept in clean water (control), exposed to the solvent (acetone control) or exposed to one of the three 17␤-trenbolone concentrations; each treatment in duplicates. The flow of the peristaltic pumps was adjusted and nominal concentrations of 17␤-trenbolone were added to the water before the eggs. The exposure period took place from 1 dpf until 60 days post hatch (dph). At 60 dph 15–20 fish from each aquarium were euthanized and prepared for histological analyses and ELISA as described in Kinnberg et al. (2007). The same day 20–25 fish from each aquarium were transferred to new aquaria with clean water for a depuration period of 170 days. The remaining fish were kept in exposure until the age of 230 dph. 2.4. Feeding and housing At the age of 3 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. Conductivity, oxygen saturation and temperature were measured twice per month. Each aquarium was aerated and the dissolved oxygen level (mean ± SEM) during the entire period was 64% ± 0.92 of the air saturation value. The fish were maintained at 26.3 ± 0.8 ◦ C (0–60 dph) and 25.4 ± 0.4 ◦ C (0–230 dph) and a 14 h light:10 h dark photoperiod. Water used during breeding, rearing and exposure was made by mixing 1 part tap water (ground water) and 3 parts deionized water and the conductivity was maintained between 220 and 250 ␮S.

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2.5. Quantification of 17ˇ-trenbolone in water samples Water samples (1 L) for chemical analysis were taken twice a month. Measurements of the actual 17␤-trenbolone concentrations were performed using solid phase extraction followed by LC–MS analysis as described by Rose et al. (2002) and Holbech et al. (2006). The average recovery of 17␤-trenbolone after solid phase extraction was 87.6% ± 3.8 (mean ± SEM) and the detection limit for trenbolone is 5 ng/L. 2.6. Histology The trunk part of each fish was embedded in paraffin after dehydration and clearing in graded ethanol/Tissue Clear series. The trunk was cut longitudinally through the entire gonadal region at 5 ␮m (60 dph) and 7 ␮m (230 dph) and afterwards stained with hematoxylin-eosin and examined by light microscopy. A minimum of ten sections from each fish was analyzed. Based on the presence of spermatogenetic cells or oocytes the gonads were classified as testis or ovaries, respectively (Fig. 1). Gonads were classified into the following stages according to Kinnberg et al. (2007) and Selman et al. (1993): Oocyte stage FI (primary growth), F2 (cortical alveolus), F3 (vitellogenic) or F4 (mature) and sperm stage M1 (sperm ducts but no spermatozoa), M2 (few/moderate spermatozoa) or M3 (abundant spermatozoa). Oocyte size, features of the nucleus, presence of cortical alveoli and vitellogenin were used to assess ovary developmental stages. As described by Holbech et al. (2006) gonads comprised of testicular tissue and only one or very few oocytes were classified as testis. Fish with a single or very few oocytes were only observed at 60 dph and as zebrafish are protogynic and hence develop provisional ovaries, these oocytes would probably undergo apoptosis if the fish were allowed to live beyond 60 dph. Only gonads comprising testicular tissue and numerous oocytes at the same time were classified as intersex gonads. By the absence of discernible germ cells the gonads were classified as undifferentiated. 2.7. ELISA Homogenization of the head and tail fraction of each fish was performed according to Kinnberg et al. (2007). Direct noncompetitive sandwich ELISA was used to quantify the vitellogenin concentration in the homogenate and the supernatant was assayed as described in Holbech et al. (2001). The method was slightly modified prior to analysis of the 230 dph fish: Dextran-HRP conjugated antibodies were replaced by a two-step process in order to enhance sensitivity of the assay. First, 150 ␮L of biotin conjugated antibody was added to each well and the plate incubated on a shaker for 1 h at room temperature. After washing five times with washing buffer 150 ␮L of streptavidin–HRP conjugated antibody was added to each well and the plate was incubated on a shaker for 1 h at room temperature. 2.8. Statistics The Bonferroni–Holm adjusted 2 test was used to examine differences in sex ratio and mortality between the control groups and the exposed groups. p < 0.05 was regarded as statistically significant. Control and acetone control groups were pooled if no significant difference was found. The sex ratio of lifetime acetone control fish was tested against the control and recovery acetone control and as no difference in sex ratio was found all control fish were pooled. Differences in body weight and vitellogenin concentration among control fish and acetone control fish were evaluated with a t test, and to meet the criteria of normality and equal variance

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Fig. 1. Light micrographs showing zebrafish sexual developmental stages. A: female stage 1 (F1); B: female stage 2 (F2); C: female stage 3 (F3); D: female stage 4 (F4); E: male stage 1 (M1); F: male stage 2 (M2); G: male stage 3 (M3); H: intersex; I: undifferentiated. Bar = 0.1 mm.

the data were log-transformed if necessary. When no significant difference between, respectively, males and females in the control groups and acetone control groups were observed, the fish were pooled into one male group and one female group. In case of a significant difference between control and acetone control the latter was used as reference. Differences in vitellogenin concentrations between the groups were evaluated by a Bonferroni–Holm adjusted one-way ANOVA or when variances were unequal the non-parametric Kruskal–Wallis one-way ANOVA. To evaluate differences in body weight also Bonferroni–Holm adjusted one-way ANOVA was used. Analyses were performed with SigmaStat® Statistical Software version 2.0. 3. Results No significant effect of the treatment was observed on the mortality during the entire period and the average survival in the aquaria was 49.2%. The average survival was based on the total number of fish remaining in the aquaria at the end of the experiment. The mortality mainly occurred during the prehatch period and larval stage. At 230 dph only one of the 298 fish was classified as intersex and no undifferentiated individuals were found. The actual concentrations of 17␤-trenbolone in the exposure groups were (mean ± SEM): 9.2(±1.48), 15.5(±1.38) and 26.2(±2.32) ng/L. Details regarding nominal and actual water concentrations of 17␤-trenbolone are presented in Supplementary Table S1 (Appendix A). 3.1. Analyses at 60 dph 3.1.1. Sex ratio and sexual development The sex ratios in all exposure groups were significantly different from the control (Fig. 2A) and 100% male populations were found at 15.5 and 26.2 ng/L. A significant reduction (p = 0.036) in the number of females in the 9.2 ng/L exposure group was observed. Overall, six fish were classified as intersex and five fish as undifferentiated and these phenotypes were not correlated to the treatment. In contrast

to the control group, no females were classified in stage F1 or F4 at 9.2 ng/L; only stage F2 and F3 females were present. In the control groups females in all four stages were observed; however, only one F4 female was found (Fig. 3A). Likewise very few males in stage 1 were observed at 9.2 ng/L. The percentage of M1 and M2 males decreased as the trenbolone concentration increased whereas the percentage of M3 males increased as the concentration increased (Fig. 3D). 3.1.2. Vitellogenin No significant difference between control fish and exposed males and females at 60 dph was observed (Fig. 4) although females exposed to 9.2 ng/L were significantly (p = 0.003) smaller compared to female control fish (see Supplementary Table S3). 3.2. Analyses at 230 dph 3.2.1. Sex ratio and sexual development 3.2.1.1. Recovery. After a depuration period of 170 days, the control sex ratio was approximately 1:1 and the intermediate and high exposed populations consisted of 100% males. As observed at 60 dph, females were present in the low exposure group and also at this time the number of females was significantly reduced (p = 0.011) (Fig. 2A and B). Loss of phenotypic male appearance was not observed in fish exposed to 15.5 and 26.2 ng/L and subsequent 170 days in clean water (Fig. 3E). After the depuration period all males exposed to 9.2 ng/L were classified in stage M3. An increasing amount of M1 and M2 males were observed as the concentration increased (Fig. 3E). All females except one were classified as F4 in the control groups and in the 9.2 ng/L groups only F4 females were observed (Fig. 3B). 3.2.1.2. Lifetime exposure. The groups exposed continuously to 15.5 and 26.2 ng/L produced 100% male populations. Females were still present in the groups receiving lifetime exposure of 9.2 ng/L (Fig. 2C). The sex ratio of the low exposure group was not significantly different (p > 0.05) from the control (Fig. 2C). Males in

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3.2.2.2. Lifetime exposure. A significant difference in weight between control males and males exposed to 15.5 ng/L (p < 0.001) and 26.2 ng/L (p = 0.049) was observed (see Supplementary Table S2). However, no difference in vitellogenin concentration was observed at any of the exposure concentrations (Fig. 6). The vitellogenin concentration of females exposed to 9.2 ng/L was significantly higher (p = 0.015) compared to the control females (Fig. 6) but no difference in weight was observed (see Supplementary Table S3). After subsampling and transferring of fish to clean water very few fish (n = 10) were left for the lifetime acetone control replicates and therefore the recovery acetone control was used when evaluating vitellogenin levels and sex ratio in the lifetime exposed fish.

4. Discussion

Fig. 2. Sex ratio in control groups and exposed groups at 60 and 230 dph. A: sex ratio at 60 dph; B: sex ratio in recovery fish at 230 dph; C: sex ratio in lifetime exposed fish at 230 dph. *Significantly different from the control groups. The number of fish per group is indicated in the bottom of each column.

all three developmental stages were present except at 15.5 ng/L, where only M3 males were found (Fig. 3F). 3.2.2. Vitellogenin 3.2.2.1. Recovery. The vitellogenin concentration of neither males nor females differed from the control fish after the depuration period (Fig. 5) and likewise no difference in body weight was observed (see Supplementary Tables S2 and S3).

To our knowledge, this is the first study to demonstrate irreversible effects on sexual development in fish exposed to environmentally relevant concentrations of an androgen. Exposure to 15.5 and 26.2 ng/L of 17␤-trenbolone during early life resulted in 100% male populations, and the masculinization was irreversible despite of a depuration period of 170 days. Recently, Zamora et al. (2008) exposed 30 days old guppy (Poecilia reticulata) to high levels of 17␤-trenbolone in the diet and ended up with an all-male population after 60 days of exposure. However, 32% of the fish were classified as intersex after a depuration period of 40 days (Zamora et al., 2008). These results might be explained by the exposure route or more possible the viviparous reproduction in guppy, where juveniles develop faster compared to egg laying species such as the zebrafish. In Australia, North America, and South America TbA is implanted in beef cattle and the vast majority of American beef cattle is treated with natural or synthetic hormones (USDA, 2000). TbA is metabolized to 17␤-trenbolone, trendione and 17␣-trenbolone in the blood circulation (Pottier et al., 1981; Schiffer et al., 2001). Those metabolites are excreted by the animals and consequently present in animal waste. The half-lives vary from ≤0.5 to 260 days as degradation is strongly affected by microbial activity, temperature and moisture conditions (Schiffer et al., 2001; Khan et al., 2008). 17␤Trenbolone is the most potent metabolite and it has a strong affinity for the human androgen receptor compared to dihydrotestosterone (Schiffer et al., 2001; Wilson et al., 2002). TbA has also entered the fitness and bodybuilding communities worldwide and it has almost replaced other anabolic steroids in recent years. The World Anti-Doping Agency has now included TbA on their list of prohibited substances. To what extent synthetic androgens constitute a risk in nature and how wildlife is affected is difficult to assess, but the effects of environmentally relevant concentrations in laboratory animals are very consistent. TbA and its metabolites have been shown to interfere with the endocrine system of several fish species leading to disruption of their sexual development (Ankley et al., 2003; Sone et al., 2005; Holbech et al., 2006). Furthermore, TbA is able to cross the placental barrier in rabbits and accumulate in fetal as well as maternal tissues (Lange et al., 2002). TbA metabolites have a low binding affinity for the sex hormone binding globulins compared to natural hormones and thus make the non-bound metabolites available for the fetus (Bauer et al., 2000). Both 17␣-trenbolone and 17␤-trenbolone have been detected in discharge from beef cattle feedlots in the United States in concentrations ranging from 10–120 ng/L and 10–20 ng/L, respectively (Durhan et al., 2006). Downstream to the discharge drain 17␤trenbolone was measured in a concentration of 7 ng/L. Therefore, wildlife may occasionally be exposed to trenbolone concentrations able to permanently alter sex ratio of laboratory animal populations as observed in the present study. Recently, an in vitro study revealed

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Fig. 3. Female (F1–F4) and male (M1–M3) sexual developmental stages in % of total females and males, respectively. A: females at 60 dph; B: females from recovery at 230 dph; C: females from lifetime exposure at 230 dph; D: males at 60 dph; E: males from recovery at 230 dph; F: males from lifetime exposure at 230 dph.

androgenic response of effluent from a feedlot, and wild populations of fathead minnows (Pimephales promelas) living downstream of the feedlot effluent outfall contained defeminized females and demasculinized males (Orlando et al., 2004). Surprisingly Sellin et al. (2009) failed to detect free 17␣-trenbolone and 17␤-trenbolone in urine and fecal slurry from steers implanted with TbA and estradiol benzoate. Neither urine nor feces from these steers induced androgenic effects in fathead minnows but estrogenic effects in the male fish were observed (Sellin et al., 2009). Steroids are generally excreted as glucoronide or sulphate conjugates, but conjugated or

organically bound steroid metabolites are probably deconjugated by microorganisms such as Escheriachia coli in water treatment plants (D’Ascenzo et al., 2003) and also desorbed from organic matter (Schiffer et al., 2001) in the environment. That might explain why Durhan et al. (2006) detected TbA metabolites in surface water. The effects of a combined estrogenic and androgenic exposure in nature are not clear. In this study effects of trenbolone were already observed at the lowest test concentration (9.2 ng/L) as the number of females was significantly reduced and at higher concentrations no females

Fig. 4. Vitellogenin concentrations in males ( ) and females ( ) at 60 dph presented as box plots with median (line within box), 25th and 75th percentiles (lower and upper boundary of box) and 10th and 90th percentiles (lower and upper whiskers). The number of male or female fish in each exposure group is indicated in the bottom of the figure.

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Fig. 5. Box plot of the vitellogenin concentrations in recovery males ( ) and females ( ) at 230 dph presented as box plots with median (line within box), 25th and 75th percentiles (lower and upper boundary of box) and 10th and 90th percentiles (lower and upper whiskers). The number of male or female fish in each exposure group is indicated in the bottom of the figure.

developed at all (Fig. 2). We found an increase in M3 males and a reduction in M1 and M2 males in all exposure groups at 60 dph which could indicate that 17␤-trenbolone accelerates the masculinization process of young males (Fig. 3D). This is consistent with a study of Orn et al. (2006). They found an increase in both testis area and the percentage of spermatozoa after exposure of male zebrafish to 50 ng/L of 17␤-trenbolone. In mosquitofish (Gambusia affins affins) premature differentiation of spermatozoa was initiated in male fry exposed to 17␤-trenbolone whereas the ovaries of ‘female’ fry contained both oocytes and spermatozoa (Sone et al., 2005). In the present study almost 60% of the control females at 60 dph were classified as F3 females. However, also F1 and F2 females were found. In the 9.2 ng/L treatment group F1 females were not observed and almost 80% were classified as F3 (Fig. 3A). This shows that only a few females at 9.2 ng/L are able to continue female development despite the presence of exogenous androgen during gonad differentiation. Since females were also found in the 9.2 ng/L groups at 230 dph the duration of the exposure period does not seem to induce additional masculinization and influence the sex ratio in the population. Hence, it is definitely not a question of exposure duration before the females turn into males, but as mentioned above some females seem to be more tolerant to the 17␤-trenbolone exposure as a few females are able to maintain female development at 9.2 ng/L (Fig. 2). However, at 15.5 ng/L female development is no longer possible, which is probably due to an imbalance of hormone levels during the sexual differentiation period (Fig. 2A–C). Our results support the irreversible masculinization observed by Fenske and Segner (2004) when they exposed

zebrafish to an aromatase inhibitor during gonad differentiation. The masculinizing effect of the aromatase inhibitor is most likely due to an accumulation of androgens during the gonad differentiation period. Ankley et al. (2003) found reduced 17␤-estradiol (E2) and testosterone plasma concentrations in female fathead minnows exposed to 0.03 and 0.1 ␮g/L of 17␤-trenbolone and at higher concentrations also males showed reduced levels of 11ketotestosterone. There might be a difference in the endogenous hormone production among the female zebrafish since some females are able to maintain female development despite of a moderate androgen exposure. Another androgen, methyltestosterone (MT), has been shown to cause both feminization and masculinization depending on the applied dose (Orn et al., 2003). This bipotential effect is ascribed to the fact that MT is aromatized by the enzyme cytochrome P450 aromatase and thereby converted into estrogens. Trenbolone is not aromatizable and therefore feminization and increased estradiol concentrations are not expected. However, 17␤-trenbolone inhibits endogenous androgen production in both male and female fathead minnows (Ankley et al., 2003) and thus the estrogen production is likely to be affected indirectly but the exact mechanism of trenbolone action has not been clarified yet (Zhang et al., 2008). Changes in sex ratio and gonad morphology have been reported to occur at TbA concentrations not affecting vitellogenin levels (Holbech et al., 2006). In agreement with those findings, we did not observe any significant changes in vitellogenin levels at any of the test concentrations except in lifetime exposed females (Figs. 4–6). However, it is quite sugges-

Fig. 6. Box plot of the vitellogenin concentrations in lifetime exposed males ( ) and females ( ) at 230 dph presented as box plots with median (line within box), 25th and 75th percentiles (lower and upper boundary of box) and 10th and 90th percentiles (lower and upper whiskers). *Significantly different compared to the control groups (p < 0.05). The number of male or female fish in each exposure group is indicated in the bottom of the figure.

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tive that female development is blocked at concentrations not able to affect vitellogenin production. As body weight could influence the vitellogenin level, differences in body weight were investigated (see Supplementary Table S2). It has been suggested that it might be possible to distinguish between ‘genetic’ females and feminized males by looking at the vitellogenin levels (Nash et al., 2004; Fenske et al., 2005). However, this parameter does not appear to give useful information after androgen exposure of juvenile zebrafish. Gross gonad morphology, fertility, fecundity and behavior are differently affected by estrogen exposure (Van den Belt et al., 2003; Brion et al., 2004; Maack and Segner, 2004; Nash et al., 2004; Fenske et al., 2005; Schafers et al., 2007) and also the estrogenic effects on zebrafish were demonstrated to be partly reversible depending on timing, exposure concentration and duration. When combining these results with our findings it is clear, that the sex determination in zebrafish is affected differently by estrogens and androgens. Nash et al. (2004) have suggested that estrogens suppress male development rather than producing functional females since some zebrafish feminized by EE2 developed testis after depuration. Ablation of primordial germ cells (PGCs) in zebrafish embryos resulted in an all-male population judged by morphological and behavioral criteria (Slanchev et al., 2005). However, the males were sterile as gonadal structures were absent in adult males. The authors conclude that PGCs are not necessary for the development of male somatic tissue with the exception of the gonad, and that the germ line is essential for both female development and accurate differentiation and survival of the gonad. Besides supporting the results from Slanchev et al. (2005) a new study has demonstrated that zebrafish without germ cells develop testis (Saito et al., 2008; Siegfried and Nusslein-Volhard, 2008). In the medaka these findings have also been supported as gonadal somatic cells of germ cell deficient medaka were demonstrated to be predisposed towards male development (Kurokawa et al., 2007). So far no sex chromosomes have been identified in zebrafish although several genes involved in male and female development have been identified (von Hofsten and Olsson, 2005; Jorgensen et al., 2008). Sexual development in zebrafish seems to be regulated by several genes located on different autosomal chromosomes and these genes should be activated at a specific time and in the right sequence in order to induce development of two separate sexes. Environmental factors are also believed to play a role in the sex determination process (Hsiao and Tsai, 2003). The present study confirms the suggestion that sex determination in zebrafish could be genetically controlled by autosomal genes. It would be interesting to know if the masculinized individuals are functional males when it comes to sexual behavior. Behavior seems to be more resistant to estrogenic exposure (Nash et al., 2004; Larsen et al., 2008, 2009) but it is not clear whether or not it applies to androgens too. The presented results also raise concern in an environmental perspective, as it is now clear that exposure to environmentally relevant concentrations of 17␤-trenbolone causes permanent masculinization in zebrafish. Acknowledgements We thank Nanna Brande-Lavridsen for help and instruction in the histology procedure. Special thanks to Bente Holbech and Vibeke Eriksen for their invaluable help in the laboratory. This project was supported by the Danish Environmental Protection Agency, Danish National Science Research Council and Nordic Council of Ministers. Appendix A. Supplementary data Supplementary data (Tables S1–S3) showing the nominal and actual water concentrations of 17␤-trenbolone and body

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