* U.S. Environmental Protection Agency, National Health and Environmental Effects Research Laboratory, Mid-Continent Ecology Division, 6201 Congdon Boulevard, Duluth, Minnesota 55804; and
Department of Anatomy and Cell Biology, School of Medicine, University of Minnesota, Duluth, Minnesota 55812
Received November 8, 2001; accepted January 14, 2002
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ABSTRACT |
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Key Words: aromatase inhibitor; endocrinology; fish; reproduction; steroids; vitellogenin.
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INTRODUCTION |
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In 1996, the U.S. Environmental Protection Agency (EPA) received a legislated mandate to develop a screening and testing program for chemicals with the potential to affect endocrine function (Food Quality Protection Act, PL#104-170; Safe Drinking Water Act, PL#104-182). A multistakeholder advisory group to the EPA suggested a number of in vitro and in vivo tests as potentially useful for identifying endocrine-disrupting chemicals (EDCs) that could affect systems controlled by estrogen, androgen, and thyroid hormones (U.S. EPA, 1998). One of the recommended tests was a short-term reproduction assay with the fathead minnow, Pimephales promelas (Ankley et al., 1998; U.S. EPA, 1998). Ankley et al. (2001) described a basic protocol for the assay: the test is initiated with sexually mature adults that are exposed to chemicals of interest for
21 days. During the exposure, data concerning reproduction and early embryonic development (fecundity, fertility, hatch) are collected. At conclusion of the test, a number of endpoints related specifically to endocrine function are assessed, including alterations in secondary sexual characteristics, gonadal condition (relative weight, histopathology), and plasma concentrations of sex steroids (E2, testosterone, 11-ketotestosterone) and vitellogenin (a precursor to egg yolk protein; Specker and Sullivan, 1994
). The assay described by Ankley et al. (2001), as well as relatively similar short-term reproduction tests with the fathead minnow, have been used to assess the effects of EDCs representative of different MOAs. The greatest emphasis to date has been upon weak and strong agonists of the estrogen receptor (Ankley et al., 2001
; Harries et al., 2000
; Kramer et al., 1998
; Miles-Richardson et al., 1999a
,b
), but there also has been some work concerning the effects of androgen receptor agonists and antagonists on fathead minnow reproduction/endocrinology (Ankley et al., 2001
; Makynen et al., 2000
). These studies have demonstrated that the assay can effectively identify and discriminate among these receptor-based MOAs. However, to support implementation of the test described by Ankley et al. (2001) as a method broadly suitable for identifying EDCs, evaluation of its performance with respect to inhibitors of steroidogenesis is critical.
The objective of this study was to evaluate the effects of the aromatase inhibitor fadrozole (4-(5,6,7,8-tetrahydroimadazo[1,5-a]- pyridin-5-yl)benzonitrile monohydrochloride; CGS 16949A) on reproductive performance and endocrinology of the fathead minnow. Fadrozole is a reversible competitive inhibitor of CYP19 that has been shown to affect E2 biosynthesis in mammals, birds, and fish (Afonso et al., 1999, 2000
; Elbrecht and Smith, 1992
; Schieweck et al., 1988
; Steele et al., 1987
). The results of this study indicate that fadrozole effectively alters steroidogenesis in the fathead minnow, producing a suite of effects that should be diagnostic for identification of CYP19 inhibitors in tests with chemicals with unknown MOAs. Our results also highlight the need for a more thorough consideration of aromatase inhibitors as EDCs of ecological concern.
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MATERIALS AND METHODS |
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Exposures were conducted using the general experimental design and techniques described by Ankley et al. (2001). Based on the results of an initial range-finding study (data not shown), the 21-day reproduction test was conducted at target fadrozole concentrations of 2, 10, and 50 µg/l. There were three replicates at each treatment level, plus a clean-water control for a total of 12 tanks. Each tank contained four female and two male adult fathead minnows from an on-site culture at the Duluth EPA laboratory. The fish were maintained at 25 ± 1°C under a 16:8 h L:D photoperiod and fed adult brine shrimp twice/day. Animals were monitored for 3 weeks prior to initiation of chemical exposure to provide tank-specific baseline data. Survival, general behavior, and fecundity were evaluated daily. A similar routine continued after starting the chemical exposure; in addition, all eggs were examined microscopically to determine fertility, and a subset of spawns from each of the treatment tanks was maintained in clean water for 5 days to determine hatching success.
After 21 days of chemical exposure, fish were removed from the test tanks and anesthetized with MS-222 (100 mg/l buffered with 200 mg NaHCO3/l), and blood was collected from the caudal artery/vein with a heparinized microhematocrit tube. Plasma was isolated by centrifugation and stored with aprotinin (0.13 units) at 80°C until determination of vitellogenin and sex steroids. Fish and gonads were weighed for determination of the gonadosomatic index (GSI), and one gonad from each fish was preserved in 1% glutaraldehyde/4% formaldehyde in 0.1 M phosphate buffer for histological analysis. Brains were removed from the fish, flash-frozen in liquid nitrogen, and stored at 80°C for subsequent determination of aromatase activity.
For histological examination of gonads, tissues were dehydrated in graded ethanol solutions and embedded in JB-4 methacrylate. Gonads from four males and eight females were examined from each of the control and the 2 µg/l treatment groups, while samples from two males and four females were evaluated at the two higher fadrozole concentrations. Longitudinally embedded gonads were sectioned at 2 to 3 µm in a step-wise fashion and stained with hematoxylin and eosin. For each ovary, three slides were made with one section from 500 µm deep into the organ and two sections from 1000 µm deep. Testes were sectioned in a similar manner, except that the sections were taken at 250 and 500 µm depths. Ovarian maturity was evaluated with respect to oocytes in the following stages: (1) primary growth, (2) cortical alveolus, (3) early vitellogenic, (4) late vitellogenic, and (5) mature/spawning oocyte (Leino and McCormick, 1997; Selman and Wallace, 1986
). In addition to assessing the overall stage of the ovary, the number of preovulatory atretic follicles per section was determined. Testicular staging conformed to Goodall et al. (1987) and Leino et al. (1990): (1) spermatocyte, (2) spermatid, (3) some sperm in lumen of seminiferous tubules, small lumina, and (4) plentiful sperm in lumen of seminiferous tubules, large lumina.
Vitellogenin concentrations were determined using an enzyme-linked immunosorbent assay (ELISA) with a polyclonal antibody to fathead minnow vitellogenin (Korte et al., 2000; Parks et al., 1999
). Plasma E2 and testosterone (T) in both sexes, and 11-ketotestosterone (KT) in males were measured using radioimmunoassay (RIA) techniques adapted to small-volume samples (Jensen et al., 2001
; U.S. EPA, 2001).
Aromatase activity is not one of the "core" endpoints in the EDC assay described by Ankley et al. (2001); however, given the expected MOA of fadrozole, activity of the enzyme in brains of male and female fathead minnows was determined in the present study. Brains were utilized as opposed to ovaries, because preliminary experiments with ovarian microsomes from individual fathead minnows failed to produce activity above background due to the small amount of recovered microsomal protein. Although there is evidence that different genes encode brain and ovarian CYP19 in fish (Chiang et al., 2001; Kishida and Callard, 2001
; Tchoudakova and Callard, 1998
), activities of both isoforms seemingly respond in a qualitatively similar manner to a variety of inhibitors (Zhao et al., 2001
). In addition, a recent study with the fathead minnow suggests significant structural homology between brain and ovarian aromatase mRNA (Halm et al., 2001
). Aromatase activity was determined by the enzymatic conversion of androstenedione to estradiol, with the release of 3H from the C-1 carbon and subsequent formation of tritiated water (Thompson and Siiteri, 1974
). A modification of the assay of Melo et al. (1999) was used. Briefly, dissected brains from individual fish (1020 mg wet weight of tissue per fish) were thawed, homogenized in 10 µl phosphate buffer (10 mM K2HPO4, 100 mM KCl, 1 mM EDTA, 1 mM DTT, pH 7.4) per mg tissue, and centrifuged at 10,000 g for 10 min. Forty to 50 µl of supernatant was incubated in phosphate buffer with 4 nM (1,4,6,7,-3H)-androstenedione (Amersham Pharmacia Biochem, Piscataway, NJ; specific activity 100 Ci/mmol) and 1 mM NADPH at 24°C for 3 h. Following the incubation, samples were placed in an ice bath, 150 µl of ethyl ether was added, and samples were held on ice for 10 min. Samples were then held at 80°C for 10 min to freeze the aqueous fraction. The ether fraction was discarded, and 300 µl of 5% dextran-coated charcoal (Sigma, St. Louis, MO) slurry was added to each tube and placed on ice for 30 min to remove any remaining steroids. Samples were centrifuged at 1500 g at 4°C for 20 min. A 300-µl aliquot of the supernatant was added to 5 ml of scintillation cocktail (Ultima Gold; Packard, Downers Grove, IL) and 3H was determined as dpm, using a Packard 2500-TR liquid scintillation counter. Data were corrected for background through analysis of samples that had been treated similarly in all respects, except they had been heated for 15 min at 90°C. Protein was determined in 5 µl samples with Bradford reagent (Sigma) and quantified by comparison to a standard curve generated with bovine serum albumin (Sigma).
Fadrozole concentrations were determined twice weekly in water from the saturator column and from each of the treatment tanks, using high-pressure liquid chromatography (HPLC) with diode-array detection (Yamagami et al., 1993). Water samples (
2 ml) were collected, placed in glass vials, and directly injected as 700-µl aliquots onto a 15-cm Adsorbosphere HS column (Alltech, Deerfield, IL) on a Hewlett-Packard 1100 HPLC with diode array detection at 230 nm. Samples were chromatographed using a gradient program with a mobile phase starting at 60% methanol/40% 50 mM phosphate buffer (pH 7.0) and increasing at 1 ml/min to 80% methanol/20% buffer. The external method of quantitation was used, with a five-point standard curve. A Lake Superior water blank and spiked sample and a duplicate from one of the exposure tanks were analyzed in conjunction with each set of test samples. Mean (SEM, n = 6) recovery of spiked samples was 103 (6.5)%, while agreement among duplicate analyses was 99 (0.34)%. The method limit of detection was about 0.5 µg/l.
For most measurement endpoints, among-treatment differences (based on tank mean values) were assessed using ANOVA followed by Dunnett's procedure (U.S. EPA, 2001). Aromatase activity was analyzed using t-tests to compare males to females within treatments (on an individual basis), and tank mean values between the control versus 50-µg/l treatment groups. When necessary, data were transformed for normalization and to reduce heterogeneity of variance. Computations were performed with Systat 7.0 (SPSS, Chicago, IL). Differences were considered significant at p 0.05.
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RESULTS |
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Brain aromatase activity was significantly reduced in both male and female fathead minnows exposed to 50 µg fadrozole/l (Fig. 1). The aromatase activity in control females was slightly higher than in control males; however, there was no significant difference in aromatase activity between the sexes within the 50-µg/l fadrozole treatment group (Fig. 1
). Measurements of brain aromatase activity were not made in the 2- and 10-µg/l treatment groups.
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DISCUSSION |
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In males, exposure to fadrozole caused a significant increase in plasma concentrations of both T and KT. The cause of this is uncertain. It seems unlikely that inhibition of CYP19 would increase androgen concentrations in males by blocking conversion of T to E2, because (1) plasma E2 concentrations in males normally are small (Jensen et al., 2001), and (2) T concentrations in females from this study were not increased despite a clear inhibition of E2 synthesis. Based on this, we speculate that the fadrozole may have inhibited some aspect of androgen degradation/excretion specific to the males. An additional observation of interest in the males was the occurrence of a relatively unique histopathology in the gonads. Specifically, there was a notable, concentration-dependent enlargement of the seminiferous tubules accompanied by an abundant accumulation of sperm in their lumina. The marked accumulation of sperm in testes at the two highest fadrozole levels could be due to enhanced sperm production related to the increase in plasma T and KT (Afonso et al., 2000
).
This study is the first with a small fish model to comprehensively assess the effects of a known aromatase inhibitor on reproductive fitness and associated endocrinology. Several of the responses observed were consistent with expectations based on studies with other classes of vertebrates (Elbrecht and Smith, 1992; Schieweck et al., 1988
; Steele et al., 1987
), including limited experimentation with fish (Afonso et al., 1999
, 2000
; Kitano et al., 2000
). Afonso et al. (1999) found that intraperitoneal injection of female coho salmon with fadrozole during late vitellogenesis (
1.5 months before spawning) significantly decreased plasma E2 concentrations, as well as ovulation (fecundity) of the fish. Injection of male salmon with fadrozole during this same period of sexual maturation advanced spermiation relative to controls; there also was some indication of increased concentrations of plasma T and KT in fish exposed to the aromatase inhibitor. Due to variations in route and timing of exposures, particularly in the context of the differential reproductive physiology of cyprinids versus salmonids, it is difficult to directly compare our results to those of Afonso et al. (1999, 2000). However, the effects of fadrozole on plasma steroid concentrations in male and female coho salmon and fathead minnows were qualitatively similar, as was the observation of reduced fecundity of females exposed to the aromatase inhibitor.
This study provides critical data relative to proposed EDC testing with the fathead minnow. First, it is clear that the suite of endpoints evaluated in the short-term reproduction assay described by Ankley et al. (2001) should effectively identify substances that inhibit aromatase activity, both as reproductive toxins and with respect to MOA. Fadrozole decreased fecundity of the fish, which was accompanied by marked alterations in gonadal histopathology in both sexes. These types of whole-animal and tissue-level alterations, while important to assessing the potential hazard of chemicals, do not directly implicate any particular toxic MOA. However, the profound reductions in circulating E2 were diagnostic of aromatase inhibition by fadrozole. This finding highlights the value of steroid determinations as a routine measurement in EDC screening/testing with fish. Because of the relatively small volumes of blood obtained from individual fathead minnows, many of the laboratories that utilize this species for EDC testing do not routinely collect steroid data. It is possible, however, through optimization of RIA techniques, to reliably measure E2, T, and KT in individual fathead minnows (Jensen et al., 2001; U.S. EPA, 2001). The ability to do so adds information not only as to the identification of MOA (as in the present study), but also is relevant to the interpretation of the potential risk of EDCs in the environment, where observations of decreased steroid concentrations in fish exposed to complex mixtures of contaminants, such as some types of effluents, are common (McMaster et al., 1996
).
Decreases in E2 were accompanied by a reduction in plasma vitellogenin concentrations, a response consistent with expectations based on basic reproductive endocrinology of oviparous vertebrates (Specker and Sullivan, 1994). There has been an emphasis on alterations in vitellogenesis as indicative of EDC exposure in fish in general and the fathead minnow in particular. However, virtually all of this attention has concerned the physiologically abnormal induction of vitellogenin in male and/or juvenile fish exposed to estrogen receptor agonists (Ankley et al., 2001
; Harries et al., 2000
; Korte et al., 2000
; Kramer et al., 1998
; Nichols et al., 1999
; Panter et al., 1998
, 2002
; Tyler et al., 1999
). Results of the present study indicate that decreases in vitellogenin concentrations in females may be just as useful for identification of EDCs that inhibit estrogen synthesis and/or signaling as induction of the protein in males is for targeting estrogen receptor agonists. Panter et al. (2002) reported that the relatively potent estrogen receptor antagonist ZM 189,154 caused small decreases of plasma vitellogenin concentrations in mixed-sex juvenile fathead minnows. Failure of the antagonist to elicit a stronger response in that study might be related to the normally small concentrations of vitellogenin (and, by inference, E2) in juvenile fish compared to that in sexually mature adult females. Based on the marked decrease in vitellogenin concentrations observed in the present study, we speculate that inhibition of vitellogenesis, either through alterations in steroid synthesis or antagonism at the level of the receptor, would be more pronounced in spawning females (actively sequestering vitellogenin in oocytes) than in juvenile fish or sexually dormant adults.
To date, most attention relative to the ecological effects of EDCs has been on compounds with the potential to act as estrogen mimics, in particular those that activate the estrogen receptor(s) (e.g., Desbrow et al., 1998; Folmar et al., 1996
; Nichols et al., 1999
; Purdom et al., 1994
). Recently, there also has been some consideration of the occurrence and effects in fish of androgen receptor agonists in effluents associated with pulp and paper mills (Bortone et al., 1989
; Jenkins et al., 2001
; Larsson et al., 2000
; Parks et al., 2001
). Little is known, however, about chemicals in the environment that might exert adverse effects through alterations in enzymes involved in steroidogenesis, including CYP19. Alterations in endocrine function and reproduction in fish exposed to complex pulp and paper mill effluents from some locations are not inconsistent with effects on steroid metabolism (McMaster et al., 1996
); however, specific chemicals and mechanisms associated with those observations have not been well defined. In a recent study, Noaksson et al. (2001) reported an association between decreased brain aromatase activity, circulating E2 concentrations, and GSI in female perch from a contaminated lake in Sweden. The results of our study with fadrozole clearly demonstrate that inhibitors of CYP19 in fish can result in significant adverse effects on reproductive fitness. This observation, coupled with more indirect evidence of inhibition of steroidogenesis in wild fish populations (e.g., McMaster et al., 1996
; Noaksson et al., 2001
), indicates the need for further consideration of aromatase inhibition as an MOA in assessing the ecological consequences of EDCs.
Controlled experimentation with single chemicals to identify causal (or correlative) relationships across different biological levels of organization can be critical to diagnostic assessments of adverse effects in wild fish populations. Three of the endpoints evaluated in this study, plasma steroid and vitellogenin concentrations and aromatase activity, have been used to varying degrees as "biomarkers" in studies focused on assessments of the possible effects of EDCs in fish from the field (e.g., McMaster et al., 1996; Noaksson et al., 2001
; Sumpter and Jobling, 1995
). There is little doubt that these types of endpoints can effectively indicate whether an animal had been exposed to a chemical(s) with a particular MOA, but these (or other) biomarkers are often questioned as to the interpretation of their biological significance in terms of fitness of the animal (Ankley et al., 1997
). Usually, it is a question of whether or not adverse effects are occurring in individuals (and populations) that is ultimately of concern to risk assessors/resource managers. For example, from an EDC perspective, there has been considerable debate as to whether observations of vitellogenin induction in male fish from the field can be correlated with adverse reproductive outcomes, or whether this response is only an indication of exposure to estrogen receptor agonists (Ankley et al., 1997
). Although the present study does not lend insight to that question, our data do indicate that, in reproductively mature female fathead minnows, decreases in (brain) aromatase activity can be associated with reduced plasma E2 and, subsequently, with decreases in vitellogenin concentrations. Further, these alterations are directly correlated, perhaps causally, with decreased fecundity of the fish. Hence, for these biomarkers (at least in the context of actively spawning females), there appears to be a reasonable basis for their utility, not only in identifying exposures to chemicals with a given MOA, but also in prediction of adverse effects at the level of the whole organism and, perhaps, through linkage to appropriate models, populations.
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ACKNOWLEDGMENTS |
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NOTES |
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1 To whom correspondence should be addressed. Fax: 218-529-5003. E-mail: ankley.gerald{at}epa.gov.
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