1Department of Zoology, 223 Bartram Hall University of Florida, Gainesville, FL 32611-8525
²Biology Department, 18952 E. Fisher Road, St. Mary's College of Maryland, St. Mary's City, MD 20686-3001
³Department of Biology University of Nebraska at Omaha, Omaha, NE 68182-0040
⁴United States Environmental Protection Agency, Research Triangle Park, NC 27711
⁵Department of Anatomy and Cell Biology Tufts University School of Medicine Boston, MA 02111
Direct correspondence to: Edward F. Orlando² Tel: (240) 895-4376 Fax: (240) 895-4996 Email: eforlando@smcm.edu 

Keywords:
concentrated animal feeding operation (CAFO), pharmaceuticals and personal care products (PPCPs), anabolic steroid hormones, H-P-G axis, environmental androgens and estrogens, aquatic ecosystem health, Pimephales promelas, gene expression.
Acknowledgements: This research was supported by a grant to EFO and LJG from the European Commission (contract # DG XII-E2/98/AF/2). In addition, the authors thank Gary Ankley, Marjorie Chow, Wendy Hessler, Martha Mann, Kyle Selcer, Charles Tyler, and Richard Stasiak for their assistance during this study.

Abbreviations: 
AR androgen receptor, CAFO concentrated animal feeding operation, Con contaminated site, DHT dihydrotestosterone, DO dissolved oxygen, E/A estrogen/androgen, E2 estradiol-17ß, E-Screen Estrogen Screen, FHM fathead minnow, FLE feedlot effluent, HW head width, IB MX isobutyl-1-methylxanthine, Int intermediate site, IO interocular, MS-222 tricaine methanesulfonate, PPCP pharmaceutical and personal care products, Ref reference site, T testosterone, USEPA United States Environmental, Protection Agency, Vtg vitellogenin

Disclaimer: 
The research described in this article has been reviewed by the National Health Environmental Effects Research Laboratory, United States Environmental Protection Agency, and approved for publication. Approval does not signify that the contents necessarily reflect the views and policies of the Agency, nor does mention of trade names or commercial products constitute endorsement or recommendation for use.

ABSTRACT

Research Over the last decade, research has examined the endocrine disrupting action of various environmental pollutants including hormones, pharmaceuticals, and surfactants in sewage treatment plant effluent. Responding to the growth of concentrated animal feeding operations and the pollutants present in their wastewater (e.g., nutrients, pharmaceuticals, and hormones), the United States Environmental Protection Agency developed a new rule that tightens the regulation of these operations. In this study, we collected wild fathead minnows (Pimephales promelas) exposed to feedlot effluent (FLE) and observed significant alterations in their reproductive biology. Male fish were demasculinized (having lower testicular testosterone synthesis, altered head morphometrics, and smaller testis size). Defeminization of females, as evidenced by a decreased estrogen/androgen ratio of in vitro steroid hormone synthesis, was also documented. We did not observe characteristics in either male or female fish indicative of exposure to environmental estrogens. Utilizing cells transfected with the human androgen receptor, we detected potent androgenic responses from the FLE. Taken together, our morphological, endocrinological, and in vitro gene activation assay data suggest two hypotheses: (1) there are potent androgenic substance(s) in the FLE, and/or (2) there is a complex mixture of androgenic and estrogenic substances that alter the hypothalamic-pituitary axis, inhibiting the release of gonadotropin-releasing hormone or gonadotropins. This is the first study demonstrating that the endocrine and reproductive systems of wild fish can be adversely affected by FLE. Future studies are needed to further investigate the effects of agricultural runoff and to identify the biologically-active agents, whether natural or pharmaceutical in origin.

INTRODUCTION

There has been a great deal of research over the last decade examining the endocrine disrupting action of various environmental pollutants (G.T. Ankley, et al. 1997, Guillette and Crain 2000). Much of this research has focused on the ability of chemical pollutants to act as estrogen or androgen receptor agonists or antagonists (McLachlan 2001). The majority of the compounds studied -- pesticides and industrial pollutants -- exhibit weak receptor affinities when compared to endogenous hormones, but can produce endocrine responses both in vitro and in vivo at environmentally relevant doses (Rooney and Guillette 2000, C.R. Tyler, et al. 1998).

Studies have begun to focus on natural hormones released from animal waste used to fertilize agricultural fields. Significant concentrations of estrogens and androgens have been reported in ponds or streams receiving runoff from fields fertilized with chicken litter (Finlay-Moore, et al. 2000, Nichols, et al. 1997, Shore, et al. 1995). In fact, depending on application rate, concentrations in runoff have been measured as high as 1,280 ng/L (Nichols, et al. 1997). Natural hormones, such as estradiol, have also been reported in ponds below cattle holding facilities and have been associated with elevated plasma concentrations of the yolk precursor protein vitellogenin in female turtles (Irwin, et al. 2001). Contamination of water systems with endogenous hormones such as estradiol-17ß and testosterone is not limited to surface waters, as estradiol has been reported in spring water from mantled karst aquifers in agricultural areas (Peterson, et al. 2000).

In addition, the presence of endogenous and pharmaceutical estrogens in sewage effluent has been studied as an example of hormonal pollution of the aquatic environment and has been reported as a factor affecting fish development and reproductive activity (Purdom, et al. 1994, C.R. Tyler, et al. 1998). Work performed below sewage treatment plants in Great Britain has documented a significant number of intersex fish when compared to rivers with less effluent (Desbrow, et al. 1996, Jobling, et al. 1998). Furthermore, these studies have reported that many males had elevated levels of estrogen-induced vitellogenin in their blood. This protein does not normally occur in males. Fractionated sewage effluent, derived mostly from domestic sources, exhibited various peaks with estrogenic activity. Those representing ethinyl-estradiol and estrone displayed the most potent estrogenic activities (Harries, et al. 1996, Harries, et al. 1997). In other countries, similar research has supported these observations and extended them by reporting that male fish exposed to sewage effluent not only have detectable plasma Vtg concentrations but also display altered plasma concentrations of testosterone and estradiol-17ß (Folmar, et al. 2000, Folmar, et al. 1996, Orlando, et al. 1999).

These studies have helped focus attention on the possible detrimental roles of pharmaceutical agents released into the environment. A wide array of pharmaceutical agents, including hormonal mimics, have been reported in sewage and open waters in various countries (Daughton and Temes 1999, Kolpin, et al. 2002, Stumpf, et al. 1999, Temes 1998). These agents include drugs commonly prescribed for the treatment of heart disease, stress, inflammation, bacterial infections (antibiotics) and birth control. Further, although veterinary drugs, such as growth promoters and antibiotics, are used extensively in agriculture, few studies have examined their presence in the environment, although some studies have recently reported the presence of these compounds in ground water near farms (Peterson, et al. 2000). Importantly, no studies have examined the possible effects of these compounds on wildlife exposed to runoff from farms using large concentrations of pharmaceutical agents, such as cattle feedlots.

In the United States, hormone supplements are used in the production of approximately 90% of the beef cattle (Baiter 1999). These supplements promote rapid growth and increase the conversion of feed to muscle mass. Currently, marketed hormone implants contain pharmaceutical grade compounds that have androgenic, estrogenic, or glucocorticogenic activities or a mixture of these activities (Schiffer, et al. 2001). The androgenic trenbolone-acetate, estrogenic zeranol, and glucocorticogenic melengestrol-acetate, are commonly used singly or combined with native steroid hormones including testosterone, estradiol-17ß, or progesterone (Schiffer, et al. 2001).

Recent studies have indicated that there is a basis for concern about the ecological effects of these pharmaceutical supplements. Trenbolone-acetate, a synthetic androgenic anabolic steroid used in cattle production, is metabolized into trenbolone-ß, the biologically active molecule, and excreted as trenbolone-a and -ß (Schiffer, et al. 2001). Trenbolone-ß has a half-life in liquid manure of more than 260 days, suggesting that it could have ecological impacts if released into the environment as runoff from the feedlots (Schiffer, et al. 2001). In another study, estrogenic activity was detected in ponds below feedlots housing a cattle herd in an academic agricultural facility (Irwin, et al. 2001).

Responding to a concern over the growth of concentrated animal feeding operations (CAFOs) and the pollutants present in their wastewater (e.g., nutrients, pharmaceuticals, hormones, etc.), the United States Environmental Protection Agency (USEPA) recently issued a new agency rule which tightens the regulation of CAFOs (USEPA 2003). The latest rule revises the existing 1976 USEPA requirements on CAFOs in two ways: (1) more CAFOs will be required to seek discharge permits under the Clean Water Act (e.g., previously exempt dry litter poultry operations) and (2) all CAFOs must develop and implement a nutrient management plan.

In our research, we examined whether endocrine activity could be detected in natural stream/river systems below feedlots by studying the reproductive endocrinology and secondary sex characteristics of wild fish populations. We examined adult fathead minnows, Pimephales promelas, living upstream and downstream of cattle feedlots in Nebraska (USA). The fathead minnow was chosen because it is a well-characterized toxicological model and native to the study region. Fathead minnows have been proposed as a sentinel species for exposure to environmental androgens and estrogens (Ankley, et al. 2001). Untreated males and female fathead minnows exposed to androgens develop increased head size and nuptial tubercles on the dorsal region of the head. Untreated females and male fathead minnows exposed to estrogens synthesize the yolk protein vitellogenin (Tyler, et al. 1999). We hypothesized that fish populations exposed to effluent from the cattle feedlots would exhibit altered sex steroid hormone titers and altered head morphology compared to fathead minnow populations from the reference site. In addition, we hypothesized that the water would contain hormonally active substances.

MATERIALS AND METHODS

Research Sites: For this initial study, we identified two affected sites: (1) a stream directly below the effluent outfall of a feedlot with a high density of penned cattle (designated the contaminated site) and (2) a stream that receives runoff from fields with dispersed cattle and agricultural activity (designated the intermediate exposure site) (Figure 1). Both sites are confluent with the Elkhorn River and have several commercial feedlots that release effluent into retaining ponds, which then drain into the river. In addition to the sites above, we identified a number of reference sites upriver from these feedlots. These streams also flowed into the Elkhorn River but with no apparent feedlot activity in the surrounding area. We were able to capture fathead minnows in sufficient numbers from only one of these sites (designated the reference site), which is located within the Oak Valley State Wildlife Management Area. At each site, water quality information was obtained that included temperature, pH, dissolved oxygen, and salinity (Table 1).

Table 1. Water quality parameters for the three sites confluent with the Elkhorn River in eastern Nebraska from which fathead minnows were collected.

Fish: During nine days in June 1999, fathead minnows (FHM, N = 97) were collected at each of the sites described above using a seine or minnow traps. Immediately upon capture, fish were placed in coolers containing aerated river water. Fish were then transferred to the University of Nebraska, Omaha where they were anesthetized with MS-222 (150 ppm, Sigma A5040, St. Louis, MO) and processed. Various morphological measurements were obtained including length (0.1 mm), mass (g), widest head width (0.1 mm), and interocular distance (0.1 mm). Hepatic tissue and gonads were removed and mass (g) obtained, then gonads were immediately transferred to an explant culture (see below). Following in vitro culturing, the gonads were fixed in neutral buffered formalin and processed for paraffin histology following standard protocol (Humason, 1997). To determine the reproductive stage of the gonad, we compared the mean values of four stages of gametogenesis in both sexes between sites (Grier, 1981; Selman and Wallace, 1989).

Gonad Cultures and Radioimmunoassays: In vitro gonadal synthesis of sex steroid hormones was examined in female and male FHM following a modification of the protocol described in the Canadian Technical Report of Fisheries and Aquatic Sciences (No. 1961) (McMaster, et al. 1995). Gonadal tissue culture media was composed of Media 199 (pH 7.4, Gibco 21200-027, Ontario, Canada), 3 -isobutyl-1-methylxanthine, (IBMX final concentration 0.1 mM, Sigma I-7018, St. Louis, Missouri USA), forskolin (final concentration 5pm, Sigma F-6886), and androstenedione (final concentration 100 ng/ml, Sigma A 9630). Culture media was sterile-filtered into an autoclaved glass bottle and stored on ice.

After gonads were excised, they were weighed, placed in glass test tubes with 1 ml of culture media, wrapped in parafilm, and incubated on a rocking plate for 6 hrs at 24°C. Parameters of the assay, including the incubation time and gonadal tissue and culture media quantity, were determined empirically from a previously conducted pilot study. Following incubation, the culture media was decanted and stored at -80°C until assayed.

In vitro production of estradiol-17ß (E₂) and testosterone (T) in female and T in male FHM were measured via radioimmunoassay on extracted culture media as described in a previously published protocol (Guillette, et al. 1995). Culture media samples were extracted twice with ethyl ether, vaporized under a stream of filtered dry air, and resuspended with 100 _µl of 0.5 M borate buffer (pH 8.0). Following resuspension of the steroid hormones, the following assay constituents were added: 200 _µl of antibody, 100 _µl of bovine serum albumin borate buffer, and 100 _µl of ³H-hormone. Final sample volume was 500 _µl and all assay tubes were run in duplicate. E₂ and T standards were also made in duplicate at concentrations of 1.56, 3.12, 6.25, 12.5, 25, 50, 100, 200, 400, and 800 pg/tube.

All samples and standards were incubated overnight at 4°C, then 500 _µl of 5% charcoal/0.5% dextran/ 0.5M phosphate buffered saline mixture were added to separate the bound from free hormone. The tubes were vortexed, centrifuged, and the supernatant containing the bound hormone was decanted. Five ml of ScintiVerse BD scintillation cocktail (Fisher Scientific, Pittsburgh, PA) was combined with the supernatant and the tubes were counted on a Beckman scintillation counter, model LS 5801 (Somerset, NJ). Extraction efficiencies of 95% for E₂ and 99% for T were used to correct raw data to actual media concentrations. Assays were validated by comparing the slopes of an internal standard curve, a media dilution curve, and the assay's standard curve. Parallelism between the internal standards, media dilutions, and assay standard curves, was confirmed using homogeneity of slopes for E₂ (p = 0.24) and T (p = 0.11) (StatView 5.0, SAS Institute, Inc., Cary, NC).

Bioassays for Hormonal Activity in Water Samples: Water was sampled in USEPA approved glass bottles concurrent with collection of the fish at the contaminated, intermediate, and reference sites. In addition, water was obtained from a retaining pond, which is located immediately at the base of the feedlot and whose outfall is the headwaters for the contaminated site. Water was refrigerated upon collection and treated with sodium azide to inhibit bacterial degradation of organic matter in samples. Samples were shipped to the laboratory of Professor Ana Soto, Tufts University for analysis of in vitro androgenic and estrogenic activity (Soto et al., in press at EHP). Additional water samples were collected one year later (June 2000), treated as stated above, and shipped to the USEPA for androgenic activity analysis (see below). Table 1 depicts information on sampling conditions and basic water quality parameters.

Preparation of Water Samples for CV-1 Androgen Receptor-Dependent Transcriptional Activation Assay: Dosing media was made using water that was obtained from the retaining pond immediately below the feedlot, as described above. Powdered DMEM (Gibco, Invitrogen Corporation, Carlsbad, CA) with 3.7 g NaHCO₃ (ICN Biochemicals, Irvine, CA) was reconstituted with 1 liter of retaining pond water and adjusted to a pH = 7.4. Media was sterile filtered (0.2 micron, Nalgene bottle- top filters, Fisher Scientific), supplemented with 5% Dextran Charcoal Serum (HyClone, Logan, UT), antibiotics added (Gibco), wrapped in aluminum foil, and stored at 4°C until use in the CV-1 transcriptional activation assay.

CV-1 Androgen Receptor-Dependent Transcriptional Activation Assay: Several experiments were conducted to determine if feedlot effluent (FLE) induced human androgen receptor (hAR) dependent gene expression in CV-1 cells (monkey kidney line, ATCC: see Parks, et al. 2001, for further description of this assay and its use in testing androgenicity of water in other aquatic systems). To determine if FLE displayed AR agonist activity, cell media was made with site water. In this experiment, 200,000 CV-1 cells were plated in a 60 mm dish and then transiently cotransfected with 50 ng pCMVhAR and 5 _µg MMTV-luciferase reporter using 5 _µl Fugene reagent in 95 _µl serum-free media (Boehringer-Mannheim, Basel, Switzerland) (seven replicate studies). Twenty-four hours after transfection, cells were dosed with 4 ml of media that was made with water from the retention pond site and incubated at 37°C with 5% CO₂. After 24 hrs of exposure, the media was removed and the cells were washed once with phosphate buffered saline then harvested with 500 _µl lysis buffer (Promega, Madison, WI). Relative light units of 0.05 ml aliquots of lysate were determined using a Monolight 2010 luminometer (Analytical Luminescence Laboratories, San Diego, CA).

Statistical Analyses: We tested for differences between sites for body length and mass, gonad mass, hormones and head morphometrics in fish by one-way ANOVA or ANCOVA (StatView 5.0) and ANOVA on the CV-1 androgen receptor-dependent transcriptional activation assays (SAS Institute, Cary, NC). If needed, data were log-transformed to obtain homogeneity of variance. Correlations between various hormones and body parameters were determined using Pearson's correlation or multiple linear regression analyses (StatView). Differences between examined groups were considered significant at P < 0.05.

RESULTS

Morphometrics: No significant difference was noted in length (P = 0.29) and mass (P = 0.70) among female fish from the three sites (Table 2). Further, no significant difference was noted in ovarian (P = 0.13) or liver (P = 0.45) mass. In contrast, interocular (IO) distance was significantly different (F = 5.6; P = 0.008) with females from the contaminated and intermediate sites having smaller distances than females from the reference site (Table 2). Head width (HW), however, was not different (P = 0.47). IO distance was correlated with HW, and the regression lines from each site have similar slopes but significantly different Y intercepts (P = 0.02), with the reference site having a higher Y value than the other two sites.

Table 2. Morphometric values (mean ± 1 SE) for female fathead minnows from three sites confluent with the Elkhorn River in eastern Nebraska*.

			 Contaminated 	Intermediate 	Reference
			 Site 		Site 		Site
Measurements 		 (N=23) 	(N=13) 		(N=19)
Length (mm) 		 5.36±0.19 	5.68±0.11 	5.68±0.12
Soma mass (g) 		 2.23±0.26 	2.48±0.15 	2.46±0.14
Gonad mass (g) 		 0.312±0.05 	0.418±.05 	0.405±.03
Liver mass (g) 		 0.065±0.008 	0.065±0.006 	0.076±0.007
Interocular distance(mm) 4.05±0.18a 	4.24±0.15a 	4.72±0.13b
Head width (mm) 	 7.07±0.24 	7.22±0.19 	7.35±0.17

*Values with different superscripts, within a row of data, are significantly different (P ≤ 0.05).
Values in rows with no superscripts are not significantly different.

As with females, no significant difference was noted in length (P = 0.14) or body mass (P = 0.15) among male fish collected at the three sites (Table 3). Male fish from all sites were significantly larger than female fish from the three study sites. A significant difference was noted in testicular (F = 4.58; P = 0.017) but not hepatic (F = 1.9; P = 0.16) mass in males (Table 3). Males from the contaminated and intermediate sites had significantly smaller testes than those from the reference site. Interocular (IO) distance was significantly different (F = 4.2; P = 0.02), with males from the contaminated and intermediate sites having reduced distances compared to males from the reference site (Table 3). Head width (HW), however, was not different (P = 0.08). IO distance is correlated with HW in males, with the regression lines from each site having similar slopes.

Table 3. Morphometric values (mean ± 1 SE) for male fathead minnows from three sites confluent with the Elkhorn River in eastern Nebraska*.

			 Contaminated 	Intermediate 	Reference
			 Site 		Site 		Site
Measurements 		 (N = 12) 	(N = 10) 	(N = 17)
Length (mm)		 6.25±0.35 	6.68±0.25 	6.85±0.07
Soma mass (g)		 3.69± 0.65 	4.06±0.46 	4.80±0.18
Gonad mass (g) 		 0.067±0.01a 	0.88±0.01a 	0.111±.01b
Liver mass (g) 		 0.107±0.02 	0.104±0.02 	0.143±0.01
Interocular distance(mm) 5.58±0.48a 	5.83±0.37a 	6.77±0.15b
Head width (mm)		 8.3±0.54 	8.64±0.39 	9.34±0.12

*Values with different superscripts, within a row of data, are significantly different (P ≤ 0.05).
Values in rows with no superscripts are not significantly different.

Histopathology: No apparent pathology was observed in any of the ovaries or testes using standard histological techniques. Also, through histological examination, we confirmed that all fish collected were adults and that the reproductive stage of the gonads in males and females did not vary among sites.

Gonadal Steroidogenesis: No significant difference in ovarian E₂ synthesis was observed among sites (P = 0.44, Fig. 2A). Ovarian mass was not correlated with E₂ synthesis (contaminated: r² = 0.074, P = 0.22; intermediate: r² = 0.115, P = 0.25; reference: r² = 0.169, P = 0.11). Mean ovarian synthesis of T was not different among sites (P = 0.08, Fig. 2B). When the data from the females were examined as an estrogen/androgen ratio (E/A), a significant difference was clearly apparent (F = 5.6; P = 0.02: Fig. 2C). Our data indicate that the females from the contaminated and intermediate sites had a defeminized sex hormone ratio, i.e., a decreased E/A ratio based on a reduction in E₂ synthesis and an increase in T synthesis (Fig. 2A, 2B).

There was a significant difference in T synthesis in vitro from testicular tissue obtained from the fish collected from the three sites (F = 5.6; P = 0.008: Fig. 3), and in vitro testosterone synthesis was lower in testes obtained from contaminated and intermediate site fish. Testosterone synthesis was not correlated with testicular weight at any of the study sites (contaminated: r² = 0.14, P = 0.21; intermediate: r² = 0.03, P = 0.61; reference: r² = 0.11, P = 0.19).

CV-1 AR-Dependent Transcriptional Activation Assays: We assessed androgenicity in seven replicate experiments (with duplicates of each replicate). Androgenicity was defined as the ability of FLE or dihydrotestosterone (DHT) to induce AR-dependent luciferase gene expression in a transfected CV-1 cell line. In every sample, and in all seven replicates, FLE induced AR-dependent luciferase gene expression. The data presented in Fig. 4 are expressed as fold induction over the control media (without FLE) and compared to the positive control of 1 nM DHT (near maximal concentration in terms of its ability to induce luciferase expression). FLE and DHT each exhibited significantly higher androgen activity than media (P < 0.0001, for each treatment versus media control). DHT- and FLE-induced responses were not significantly different (P = 0.35) from each other.

DISCUSSION

To our knowledge, this is the first study to document endocrine disruption in fish exposed to feedlot effluent (FLE). Wild fish collected below a feedlot exhibited altered reproductive biology including decreased T synthesis, altered head morphometrics, and smaller testis size in males and decreased E/A ratio in female fish. We did not observe overt characteristics in either male or female fish suggesting environmental exposure to estrogens. With an in vitro assay utilizing cells transfected with the human androgen receptor, we detected potent androgenic responses from the FLE. Taken together, our morphological, endocrinological, and in vitro gene activation assay data suggest two hypotheses: (1) there is an androgenic substance(s) in the FLE and/or (2) there is a mixture of endocrine active substances that alter the hypothalamic-pituitary-gonadal axis. Work performed in collaboration with the laboratory of Professor Ana Soto adds further support to the hypothesis that androgens are present in the FLE, as her lab observed androgenic activity. However, her lab also report estrogenic activity in FLE using the MCF-7 cell in vitro E-Screen assay, suggesting that there could be a complex mixture of natural and pharmaceutical compounds in the effluent (Soto et al., in press at EHP).

Our data clearly demonstrate androgenic activity from water obtained below feedlots. However, it does not identify the causal agents. Androgenic activity could be due to natural androgens found in fecal material or androgenic pharmaceuticals used in growth implants (Meyer 2001). Natural androgens have relatively short half-lives in feces and in the open water of retaining ponds (Meyer 2001). In contrast, recent studies demonstrate that metabolites of synthetic androgens (such as trenbolone-ß from trenbolone-acetate) used in growth implants have longer half-lives. Approximately 27.5% of the initial concentration of trenbolone-ß was still present in manure piles 4.5 months after deposition (Schiffer, et al. 2001). Natural steroids appear to be rapidly degraded, with half-lives measured on the order of days to hours. No literature could be found regarding the relative persistence of zeranol or melengestrol in feedlot retaining ponds, however.

Trenbolone-ß acts as a potent androgen agonist in the CV-1 cell assay used to test FLE in this study (Wilson, et al. 2002). In fact, its potency was equal or greater to that of the positive control, DHT, at similar concentrations. Trenbolone-acetate is known to be 8-10x more potent than native testosterone in cattle (Schiffer, et al. 2001). Furthermore, in an in utero screening assay, maternal trenbolone-ß increased ano-genital distance and attenuated the display of nipples in female rat offspring (Wilson, et al. 2002).

In a recent laboratory study, fathead minnows were exposed to trenbolone-ß. The reproductive biology of female and male FHMs was severely altered from this exposure (Ankley, et al. 2003). In females, fecundity decreased, male-like secondary sex characteristics developed (nuptial tubercles) and plasma concentrations of T, E₂, and vitellogenin were all significantly decreased. In male FHMs, plasma concentrations of 11-ketotestosterone were decreased and E₂ and vitellogenin were increased. Although difficult to compare directly due to differences in experimental design, data from our field study support the results of this laboratory study.

Trenbolone-ß binds the FHM androgen receptor with greater affinity than T (Ankley, et al. 2003). In male FHMs, trenbolone-ß could act at the level of the hypothalamus or pituitary to depress GnRH and/or GtH synthesis and/or release, leading to decreased T synthesis, testicular mass, and interocular distance as seen in the contaminated site males. Female FHMs exposed to FLE at the intermediate and contaminated sites had decreased E/A ratios that stemmed from a decrease in ovarian E₂ and an increase in T synthesis during in vitro culture. That is, if the hormones were examined individually, no significant difference was observed among sites. However, when a ratio was calculated, it was obvious that ovarian steroidogenesis was altered in fish obtained from the intermediate and contaminated sites. This result suggests that some component of the FLE has the potential to inhibit ovarian aromatase, the enzyme which converts T to E₂ (Norris 1997). Interestingly, trenbolone-ß at certain concentrations has been shown to weakly bind the FHM estrogen receptor, induce vitellogenesis in male FHMs, and weakly bind the rainbow trout estrogen receptor in an in vitro transfected yeast system (Ankley, et al. 2003, Le Guevel and Pakdel 2001). Future research should investigate what constituent(s) of the FLE may be inhibiting aromatase synthesis or action.

Other compounds that are strong anabolic agents, such as the mycotoxin zearalanol, are estrogenic in cattle, humans, rainbow trout (Oncorhynchus mykiss), and Atlantic salmon (Salmo salar) (Arukwe, et al. 1999, Le Guevel and Pakdel 2001). Zearalanol is also known to depress FSH and LH concentrations in cattle. Zearalanol, measured as resorcylic acid lactones, was not detectable in the study by Soto et al. (Soto et al., in press at EHP). Furthermore, we do not know, presently, if zearalanol can interact with GnRH or GtH receptors in fish.

Water quality parameters obtained during this study suggested that the responses observed in fish were unlikely to be complicated by differences in the aquatic environment (Table 1). No fish were found in the retaining pond immediately below the feedlot. This site had very low dissolved oxygen (DO) levels (0.7 ppt) and relatively high salinity (1.2 ppt). When the contaminated sites (where fish were obtained) were compared to the reference site, it was apparent that DO was slightly different, as was salinity. The slightly lower observed DO is not surprising given the eutrophic nature of the effluent-laden streams where fish were caught. Salinity was also elevated at the contaminated site versus the other sites, but the levels reported here should have little effect on the fish, as the differences were less than 1 ppt. Thus, it is unlikely that these variables significantly influenced the endpoints measured in this study.

We were not able to identify sites (feedlots) where only endogenous fecal steroids would be in the runoff. That is, all the feedlots we identified used growth implants in their cattle. We had hoped to identify sites that had operations that raise cattle without hormone supplements and searched extensively for such locations in the same region. All of the operations we identified, which did not use hormone implants, did not raise cattle in a feedlot setting. These implant-free cattle are usually free-ranging cattle, i.e., they are raised at low density on open rangelands. Future studies are needed to examine fish exposed to slurries of manure from treated and untreated animals. Given the recent publication documenting wide-scale contamination of US water bodies with numerous pharmaceutical agents (Kolpin, et al. 2002), future work -- such as that presented in this study combined with intensive environmental chemistry -- is urgently needed if we are to understand the possible adverse effects of these compounds on aquatic ecosystem health.

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