Effects of the Feedlot Contaminant 17α-Trenbolone on Reproductive

Environ. Sci. Technol. 2006, 40, 3112-3117

Effects of the Feedlot Contaminant 17r-Trenbolone on Reproductive Endocrinology of the Fathead Minnow KATHLEEN M. JENSEN,* ELIZABETH A. MAKYNEN, MICHAEL D. KAHL, AND GERALD T. ANKLEY U.S. Environmental Protection Agency, Office of Research and Development, National Health and Environmental Effects Research Laboratory, Mid-Continent Ecology Division, 6201 Congdon Boulevard, Duluth, Minnesota 55804

Trenbolone acetate is a growth promoter widely used for beef production in the U.S. Two biologically active metabolites of the acetate, 17β- and 17R-trenbolone, are ligands of vertebrate androgen receptors and comparatively stable in the waste of treated animals. Both have been detected in surface water associated with beef feedlots, suggesting a potential risk to aquatic animals. In previous work we evaluated the effects of β-trenbolone on reproductive endocrinology of the fathead minnow (Pimephales promelas) in a 21-day test. The purpose of the present study was to conduct a similar set of experiments with R-trenbolone which, based on binding to mammalian androgen receptors, was expected to be less potent than β-trenbolone. Fecundity of the fish was significantly reduced by R-trenbolone with an EC50 (95% confidence interval) of 0.011 (0.007-0.016) µg/L. In females, R-trenbolone reduced plasma vitellogenin and steroid concentrations and also induced the production of dorsal nuptial tubercles, structures normally present only in spawning males. Overall, effects of R-trenbolone on the reproductive system of the fish were qualitatively and quantitatively quite similar to those caused by β-trenbolone. Part of this similarity might arise from the fact that a substantial amount of the R-trenbolone appeared to be converted to β-trenbolone by the fish. Tissue concentrations of the β-isomer were consistently similar to or greater than concentrations of R-trenbolone, despite the fact that no β-trenbolone was detected in the exposure water. The present study demonstrates the importance of considering both R- and β-trenbolone in assessing the potential ecological risk of androgens associated with beef feedlot discharges.

Introduction Environmental contaminants that adversely affect reproduction and development through alterations in endocrine function in humans and wildlife have been identified as an issue of global concern (1). Fish, clearly have been affected by endocrine-disrupting substances in the environment (2, 3). Initial studies in the UK, and later elsewhere, demonstrated the occurrence of estrogen receptor agonists, consisting of * Corresponding author [email protected].

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natural hormones and xenobiotic chemicals, that feminized both feral and caged fish (4-8). Although estrogenic chemicals are important, there is evidence that environmental contaminants which interact with the androgen receptor also may be adversely affecting wildlife (9). For example, masculinization of fish exposed to discharges from pulp and paper mill effluents and runoff from beef feedlots has been reported (10-12), and been associated with in vitro androgenic activity of water samples from affected sites (12-14). The identity of causative androgen agonists in pulp and paper mill effluents currently is uncertain (15, 16); however, different lines of evidence suggest that steroidal chemicals could contribute to androgenic activity of feedlot discharges. In addition to the natural steroids produced (and excreted) by livestock, a variety of steroids are used to enhance growth of animals (17). For example, the majority of beef production in the U.S. utilizes trenbolone acetate as an anabolic steroid to promote production of muscle mass (17, 18). Trenbolone acetate is administered to cattle via implants which release the acetate to the blood where it is hydrolyzed to 17βtrenbolone (19), a potent androgen receptor agonist in mammals (20). An additional metabolic product of trenbolone acetate is 17R-trenbolone, which predominates over the β-isomer by a ratio of about 10:1 in wastes excreted by treated livestock (19). Both metabolites are quite stable in the environment with half-lives of about 260 days in liquid manure (19). This has led to speculation that trenbolone metabolites could contribute to androgenic activity associated with feedlot runoff (12), and might cause adverse effects in fish. To help address the latter possibility, Ankley et al. (21) conducted a 21-day reproduction test with the fathead minnow (Pimephales promelas) exposed to β-trenbolone. They found that a concentration of about 0.027 µg/L of the androgen caused female fathead minnows to develop male secondary sexual characteristics (dorsal nuptial tubercles) and significantly decreased fecundity (egg production). In another study, a 28-day exposure to β-trenbolone (1-10 µg/ L) induced the formation of a gonopodium-like structure on the anal fin of adult female mosquitofish (22). Although trenbolone acetate has been used as a livestock growth promoter for many years (17), little is known concerning concentrations of trenbolone metabolites in natural environments. In a recent study we documented the occurrence of both compounds in water with in vitro androgenic activity that had been collected from a beef feedlot discharge where trenbolone acetate implants had been used (23). Measured concentrations of R- and β-trenbolone in samples from the site varied temporally but both were present, at least occasionally, at concentrations greater than 0.01 µg/L (23). Based on toxicity data collected for β-trenbolone (21), these concentrations could be of potential concern with respect to adverse effects on fish. There are no toxicity data, however, for R-trenbolone in any fish species. The objective of the present study, therefore, was to evaluate the reproductive toxicity of R-trenbolone in the fathead minnow with the 21-day reproduction test protocol that had been used for β-trenbolone, as a basis to better understand the potential risk of the two androgens in aquatic systems.

Materials and Methods In this paper we report the results from two experiments with R-trenbolone. In the first study, target water concentrations of the androgen were 0 (control), 0.175, 0.7, 1.75, 3.5, and 7.0 µg/L. These treatments were chosen based on (a) concentrations of β-trenbolone that produced effects in the study by Ankley et al. (21), in conjunction with (b) a potency 10.1021/es052174s Not subject to U.S. copyright. Publ. 2006 Am. Chem.Soc. Published on Web 03/31/2006

“adjustment” factor of 10 derived from competitive binding studies with mammalian androgen receptor(s) indicating approximately a 10-fold higher affinity for β-trenbolone than R-trenbolone (24, 25). However, in the first study, no reproduction was observed in the fish at any of the R-trenbolone concentrations. Hence, a second experiment was performed at lower target water concentrations of 0, 0.003, 0.01, 0.03, and 0.1 µg/L. Toxicity tests were conducted using the basic protocol described by Ankley et al. (26). Solvent-free stock solutions of R-trenbolone (Aventis, France, >99.6% purity; or Hayashi Pure Chemical Industries, Osaka, Japan, 99.9% purity) were prepared by dissolving neat chemical in water to a concentration of 72.8 (first experiment) or 6.75 (second experiment) mg/L. The stock solutions were subsequently diluted with Lake Superior (control) water to achieve the test concentrations indicated above, and delivered to the exposure tanks at a flow rate of approximately 45 mL/min. Sexually mature fathead minnows (5-6 months old) from an on-site culture were held in the test system for at least two weeks prior to initiation of the chemical exposure, during which fecundity was evaluated daily. The fish were held at 25 ( 1 °C under an 16:8 light:dark photoperiod, and fed adult brine shrimp (Artemia) twice daily. In the first study, duplicate exposures were conducted at each test concentration using a group-spawning approach with four females and two males per tank. In the second study, a paired-spawning design was used, with eight pairs of fish at each treatment level. Exposures were conducted for 21 days under the same conditions as the acclimation phase of the test, except that fertility and hatching success also were assessed. At the conclusion of the exposure, fish were anaesthetized with buffered tricaine methanesulfonate (MS-222, 100 mg/L with 200 mg NaHCO3/L), and the occurrence and degree of expression of nuptial tubercles was assessed (27). Blood was collected from the caudal vein/ artery of the fish, and plasma was separated by centrifugation and stored at -80 °C until measurement of steroids and vitellogenin. Sex steroids (17β-estradiol, testosterone, 11ketotestosterone) were determined using a radioimmunoassay technique adapted to small volumes (27), and vitellogenin was measured using an enzyme-linked immunosorbent assay with a fathead minnow polyclonal antibody (28, 29). Concentrations of R-trenbolone were determined in stock solutions and water samples from each exposure tank at least once weekly during both 21-day assays. Samples from the first assay were analyzed by direct injection of 500 µL of test water onto an Alltech Adsorbosphere HS column (Deerfield, IL) on an Agilent 1100 high-pressure liquid chromatograph (HPLC) with fluorescence detection at excitation and emission wavelengths of 359 and 458 nm, respectively. A gradient program starting with 50% methanol in water increasing to 80% methanol with a flow rate of 0.9 mL/min and column temperature of 35 °C was used. The analyte was quantified using an external standard method. In the first experiment, the limit of detection was 0.12 µg/L. The mean (standard deviation [SD]) recovery of matrix-spiked samples was 91% (15.8%, n ) 6). Agreement among duplicate samples was 99% (2.1%, n ) 5). In the second study, water from the nominal 0.1 µg/L test concentration also was analyzed using direct injection similar to the first study. To increase the method limit of detection, water (25-100 mL) from the three lower test concentrations in the second experiment was concentrated on 3 mL C18 columns (J. T. Baker, Phillipsburg, NJ), which subsequently were eluted with 2 mL of 100% methanol. Extracts were evaporated to dryness, and 1 mL of 10% methanol was used to resuspend samples before conducting HPLC analyses under the same conditions as used for the direct injection samples. The mean (SD)

recovery of spiked samples analyzed when using the C18 concentration step was 75% (8.2%, n ) 6). Agreement among duplicate samples was 91% (3.0%, n ) 6). The limit of detection was 0.0004 µg/L for the concentrated samples. In both experiments, the possible presence of β-trenbolone was monitored in conjunction with the R-trenbolone analyses; in no instance was the β isomer detected in the exposure water (limit of detection 0.0004 µg/L). After blood was removed from the fish, they were wrapped in foil, placed in plastic zip-lock bags and frozen at -20 °C until extraction. Two fish of each sex from each of the treatment concentrations in the first experiment were individually homogenized with 15 mL of acetonitrile for 1-2 min. The extract was transferred to a polypropylene centrifuge tube and centrifuged at 3000g for 20 min at -10 °C. The supernatant was evaporated under nitrogen to about 10 mL, placed into a freezer at -20 °C for 1 h, transferred to clean tubes, and then further evaporated to 5 mL. Extracts were diluted with equal portions of HPLC-grade water before spinfiltering (using Alltech micro-spin filter tubes), and 100 µL aliquots were injected onto the HPLC under the same conditions as used for the water samples. In the first experiment, the detection limit for R- and β-trenbolone in the fish tissues ranged from 0.91 µg/kg in the largest fish to 3.25 µg/kg in the smallest fish. The mean (SD) recovery of R-trenbolone in spiked samples was 93% (16.3%, n ) 3), while the recovery for β-trenbolone was 102.9% (14.4%, n ) 2). In the second study, fish from the 0.1 µg/L treatment group were extracted in the same fashion, but with additional cleanup and concentration steps to enhance analytical sensitivity. To further increase sensitivity in the second study, some of the analyses were conducted with composite samples (pooled by sex) rather than individual fish. After the freezing step in the preparation described above, extracts of fish from the second experiment were allowed to warm to room temperature, 50 µL of 10 mM phosphate buffer (pH 3) was added to each 10 mL extract, and the samples were passed through a 600 mg Varian Bond Elut LRC Accucat Multimode column (Harbor City, CA). Ethyl acetate eluates were evaporated to dryness, dissolved in 2 mL of acetone:hexane (10:90) and subjected to further cleanup using 500 mg Waters silica (Sep-Pak) columns (Milford, MA). Samples were loaded onto the silica columns and washed with 10 mL of the 10:90 acetone:hexane. The columns were then eluted with a 15:85 acetone:hexane mixture. The extracts were evaporated to dryness using nitrogen, and 1 mL of 30% methanol in HPLCgrade water was added before HPLC analysis of 200 µL sample aliquots. The HPLC conditions were as described above, except that the excitation and emission wavelengths were shifted slightly to 364 and 460 nm, respectively, to minimize interferences. The detection limit for R- and β-trenbolone was 0.02 µg/kg for the tissue samples from the second experiment. The mean (SD) recovery of R-trenbolone in spiked samples was 63.0% (33.4%, n ) 4), while recovery of β-trenbolone was 38.8% (25.0%, n ) 4). Differences between treatments at the conclusion of the test were assessed with ANOVA followed by Dunnett’s procedure. Concentrations of trenbolone resulting in a 50% reduction in fecundity (EC50) were estimated using logistic regression (30). When necessary, data were transformed (log or square root) for normalization and/or to reduce variance heterogeneity. Analyses were performed using Systat 9 (SPSS, Chicago, IL). Results were considered significant at P e 0.05.

Results Table 1 summarizes exposure data for the two studies. Water concentrations of R-trenbolone in the test tanks were comparatively stable, and always within 75% of target concentrations for the first study, and within 90% of nominal VOL. 40, NO. 9, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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TABLE 1. Water Concentrations of 17r-Trenbolone and Tissue Concentrations of 17r-Trenbolone and 17β-Trenbolone in Fathead Minnows from Two 21-Day Studies tissue concentration (µg/kg) target water concentration (µg/L)

17r-trenbolone

17β-trenbolone female

measured water concentration (µg/L)

male

female

male

0 (control) 0.175 0.70 1.75 3.5 7.0

BDLa 0.13 ( 0.016c 0.66 ( 0.40 1.3 ( 0.27 2.8 ( 0.19 7.1 ( 0.27

Study 1 BDLb BDL 3.2 ( 0.62d 10 ( 4.0 8.6 ( 2.3 17 ( 0.9

BDL BDL 3.4 13 ( 4.6 11 ( 4.4 22 ( 2.8

BDL BDL 4.6 ( 2.3 12 ( 3.8 24 ( 2.0 20 ( 4.4

BDL BDL 4.1 26 ( 12 18 ( 11 42 ( 14

0 (control) 0.003 0.01 0.03 0.10

BDLa 0.0035 ( 0.0008c 0.0097 ( 0.0008 0.032 ( 0.0031 0.094 ( 0.0033

Study 2 BDLb e NA NA NA 0.20 ( 0.18f

BDL NA NA NA 0.37 ( 0.21

BDL NA NA NA 0.12 ( 0.03

BDL NA NA NA 0.41 ( 0.25

a Below detection limit (0.12 µg/L for Study 1 and 0.0004 µg/L for Study 2). b BDL (0.91 µg/kg in largest fish and 3.25 µg/kg in smallest fish from Study 1; 0.02 µg/kg for pooled fish from Study 2). c Mean ( SD (n ) 6 for Study 1, and n ) 3 for Study 2). d Mean ( SD (n ) 2 except for females from the 0.70 treatment where n ) 1). e Not analyzed; values were assumed to be BDL. f For R-trenbolone: n ) 4 males; two individuals and two composites comprised of two fish each, and n ) 4 females; three individuals and one composite comprised of two fish. For β-trenbolone: n ) 2 males; two composites comprised of two fish each, and n ) 3 females; individuals.

concentrations for the second experiment. Residues of R-trenbolone were detected in whole-fish extracts from the four highest treatment groups in the first study (Table 1). Tissue concentrations in animals from the 0.7 µg/L treatment were the lowest, while the highest concentrations were in animals from the 7 µg/L treatment. Concentrations of R-trenbolone in fish from the 1.75 and 3.5 µg/L treatments were intermediate and similar in value (Table 1). Somewhat surprisingly, β-trenbolone exhibited a similar trend in the exposed animals, at concentrations slightly greater than those of R-trenbolone (Table 1). In the second experiment, as in the first, both R- and β-trenbolone were detected in animals exposed to 0.1 µg R-trenbolone/L (Table 1). In both experiments there was a tendency for females to have slightly greater tissue concentrations of trenbolone than males; however, sample sizes were too small to discern whether sex-related differences in trenbolone tissue concentrations truly existed. There was no treatment-related mortality of fish exposed to R-trenbolone in either study. In the first experiment, control animals reproduced normally over the 21-day assay; however, there was virtually no egg production in any of the R-trenbolone treatments (data not shown). At the conclusion of 21 days, nuptial tubercles were not present in controls, but were observed in females from all of the R-trenbolone treatments, with the most pronounced development of male secondary sex characteristics in females from the 7 µg/L test concentration (data not shown). Based on the effects on fecundity and tubercle induction, R-trenbolone appeared to be a more potent androgen in the fathead minnow than we had expected, so additional testing at lower concentrations was needed to define effect/no-effect thresholds. Given this, additional diagnostic measurements of effects on endocrine function (i.e., plasma steroids and vitellogenin) were not conducted with samples from the first study. In the second experiment, exposure to R-trenbolone significantly reduced egg production in both a time- and concentration-dependent fashion, with an EC50 (95% confidence interval) of 0.011 (0.007-0.016) µg/L based on cumulative fecundity over the final two weeks of the test (Figure 1). Neither fertility (>98% for all treatments) nor hatching success (>92% for all treatments) were affected by R-trenbolone exposure. In the second study, R-trenbolone caused female fathead minnows to develop male-like nuptial tubercles in 2 of 8 and 5 of 8 animals, respectively, from the 0.03 and 0.1 µg/L 3114

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FIGURE 1. Effects of 17r-trenbolone on fecundity of the fathead minnow during a 21-day test. treatments (Figure 2a). Plasma vitellogenin in females from the two highest treatment groups (Figure 2b), as well as plasma estradiol and testosterone concentrations (Figure 2c,d) were significantly reduced by R-trenbolone exposure. There were no marked effects of R-trenbolone on any of the variables measured in the males, including tubercle score (Figure 3a), plasma concentrations of vitellogenin (Figure 3b), estradiol (Figure 3c), or testosterone and 11-ketotestosterone (Figure 3d).

Discussion Implants containing trenbolone acetate are widely used in the U.S. and elsewhere to enhance growth of cattle (17, 18). The acetate is hydrolyzed to 17β-trenbolone, a high-affinity ligand for the mammalian androgen receptor which, ostensibly, is primarily responsible for the anabolic properties of the implants (20). A second metabolite, 17R-trenbolone, is formed via epimerization of the β-trenbolone, and both isomers are excreted by treated livestock (19). Concentrations of R-trenbolone have been reported to be about 10-fold higher in animal waste than β-trenbolone, but both are quite stable with half-lives of around 260 days (19). Past data from competitive binding assays with mammalian androgen receptors suggest that R-trenbolone would be expected to be about an order of magnitude less potent than β-trenbolone (24, 25). Based on this, our initial experiment with R-trenbolone employed a water concentration series about 10-fold

FIGURE 2. Effects of a 21-day exposure to 17r-trenbolone on female fathead minnow: (a) nuptial tubercle score, and plasma concentrations of (b) vitellogenin, (c) 17β-estradiol, and (d) testosterone. Data are expressed as mean ( SE with sample size in parentheses. Asterisks indicate significant differences from the control.

FIGURE 3. Effects of a 21-day exposure to 17r-trenbolone on male fathead minnow: (a) nuptial tubercle score, and plasma concentrations of (b) vitellogenin, (c) 17β-estradiol, and (d) the androgens testosterone (9) and 11-ketotestosterone (0). Data are expressed as mean ( SE with sample size in parentheses.

higher than that used in a previous 21-day fathead minnow study with β-trenbolone, where the lowest-observable effect concentration was 0.027 µg/L for reduced fecundity and masculinization of females (21). However, in the first R-trenbolone study, the fish did not spawn at the lowest concentration tested (0.175 µg/L), and there was evidence of masculinization of the females at this, as well as the greater concentrations of the androgen. In a second experiment, R-trenbolone decreased fecundity (and masculinized females) at test concentrations below 0.1 µg/L, so based on water concentrations of the two chemicals, R-trenbolone clearly is not 10-fold less potent than β-trenbolone. Because of differences in the concentration ranges tested in our definitive β-trenbolone study (21) versus the second R-trenbolone experiment described in this paper, it is difficult to precisely compare potencies of the two androgens for all of the endpoints evaluated in the two studies. However, if one considers the EC50 for fecundity inhibition for R-trenbolone from the present study (i.e., 0.011 µg/L) versus that for β-trenbolone (0.018 [0.009-0.037], calculated using data from Ankley et al. [21]), the potency of the two chemicals appears to be quite comparable. In addition to having relatively similar potencies for effects on fecundity and masculinization of females, the pattern of biochemical responses in the fish to the two trenbolone isomers was quite similar. For example, in females, β-trenbolone caused significant decreases in plasma vitellogenin, estradiol, and testosterone (21), a pattern identical to that observed in our present study with R-trenbolone. Also similar to β-trenbolone was the comparative lack of responsiveness of any of the endpoints to R-trenbolone in males. Given these observations, it seems reasonable to assume that the two

chemicals are operating via the same toxic mechanism of action in the fish. There are two possible explanations for the somewhat unanticipated degree of reproductive/endocrine toxicity of R-trenbolone. First, it could be that the fathead minnow androgen receptor does not have a greater affinity for βthan R-trenbolone. In support of this, Wilson et al. (31) reported that the two trenbolone isomers had comparable competitive binding affinities for a cloned fathead minnow androgen receptor transiently transfected into COS-1 cells. An alternative explanation for the seeming potency of R-trenbolone is that it is converted to β-trenbolone by the fish, which then may be predominantly responsible for observed effects on endocrine function. This possibility is supported by our observation that, in both studies, greater tissue concentrations of β- than R-trenbolone were observed in animals exposed to R-trenbolone via the water. So, evidence exists that either (or both) explanation(s) could account for the roughly equivalent toxicity and pattern of responses of R- and β-trenbolone in the 21-day fathead minnow reproduction assay. Although it is impossible at present to identify which of the scenarios explains the relative potency of R- versus β-trenbolone, this is perhaps not all that critical from a pragmatic perspective of making an initial assessment of the potential ecological risk of mixtures of the two androgens. Durhan et al. (23) conducted a study in which water samples were collected over a period of several months from a beef feedlot where trenbolone acetate implants were used. There were marked temporal variations in the samples; however, those collected from a direct discharge from the VOL. 40, NO. 9, 2006 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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feedlot displayed significant androgenic activity in a (CV1) cell line transiently transfected with a human androgen receptor-reporter gene construct (14). Measurements of both R- and β-trenbolone also were made in the samples from nine different collection periods. As was the case for androgenic activity, there was significant temporal variability in the samples; however, measurements made with the same HPLC method used in the present study indicated the occurrence of both metabolites in the discharge samples. Six of the nine samples had detectable R-trenbolone, with concentrations ranging from less than 0.01 to about 0.12 µg/L. Two of the nine samples had detectable β-trenbolone at concentrations of about 0.01 and 0.02 µg/L. More intensive confirmatory analyses via gas chromatography-mass spectroscopy of one of the samples containing both metabolites (based on the HPLC analysis) indicated R- and β-trenbolone concentrations of 0.008 and 0.012 µg/L, respectively. Assuming roughly equipotent and additive toxicity (based on a common mechanism of action) for the two metabolites, concentrations detected in the Durhan et al. (23) study clearly are within a range that could adversely affect reproduction of the fathead minnow (21, present study). Further, these effects at the individual level could easily be of sufficient magnitude to result in population-level impacts in this species (32). In summary, R-trenbolone, like β-trenbolone, is a potent reproductive endocrine toxicant in the fathead minnow. To our knowledge, this is the first report in the open literature of the reproductive toxicity of R-trenbolone in a nonmammalian species. Given the wide use of trenbolone acetate in cattle feeding operations, the seeming persistence of R- and β-trenbolone in the environment, and the fact that the R-isomer may occur in far greater quantities than β-trenbolone, further studies concerning the occurrence and potential effects of R-trenbolone are warranted.

Acknowledgments We thank Elizabeth Durhan and Vickie Wilson for providing valuable comments on an earlier draft of this paper. Discussions with Russ Erickson greatly aided statistical analysis of test data. Roger LePage and Diane Spehar assisted in manuscript preparation. The research described in this document was funded wholly by the U.S. Environmental Protection Agency. It has been subjected to review by the National Health and Environmental Effects Research Laboratory and approved for publication. Approval does not signify that the contents reflect the views of the Agency, nor does mention of trade names or commercial products constitute endorsement or recommendation for use.

Literature Cited (1) World Health Organization. Global Assessment of the State-ofthe-Science of Endocrine Disruptors; International Programme on Chemical Safety: Geneva, Switzerland, 2002. (2) Ankley, G. T.; Giesy, J. Endocrine disruptors in wildlife: A weight of evidence perspective. In Principles and Processes for Evaluating Endocrine Disruption in Wildlife; Kendall, R., Dickerson, R., Giesy, J. P., Suk, W., Eds.; SETAC Press: Pensacola, FL, 1998; pp 349-367. (3) Tyler, C. R.; Jobling, S.; Sumpter, J. P. Endocrine disruption in wildlife: A critical review of the evidence. Crit. Rev. Toxicol. 1998, 28, 319-361. (4) Purdom, C. E.; Hardiman, P. A.; Bye, V. J.; Eno, N. C.; Tyler, C. R.; Sumpter, J. P. Estrogenic effects of effluents from sewage treatment works. Chem. Ecol. 1994, 8, 275-285. (5) Folmar, L. C.; Denslow, N. D.; Rao, V.; Chow, M.; Crain, D. A.; Enblom, J.; Marcino, J.; Guillette, L. J., Jr. Vitellogenin induction and reduced serum testosterone concentrations in feral male carp (Cyprinus carpio) captured near a major metropolitan sewage treatment plant. Environ. Health Perspect. 1996, 104, 1096-1101. (6) Desbrow, C.; Routledge, E. J.; Brighty, G. C.; Sumpter, J. P.; Waldock, M. Identification of estrogenic chemicals in STW 3116

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Received for review October 31, 2005. Revised manuscript received March 1, 2006. Accepted March 7, 2006. ES052174S

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