Toxicological Sciences 68, 389-402 (2002)
Copyright © 2002 by the Society of Toxicology
ENVIRONMENTAL TOXICOLOGY |
Transgenerational and Developmental Exposure of Japanese Medaka (Oryzias latipes) to Ethinylestradiol Results in Endocrine and Reproductive Differences in the Response to Ethinylestradiol as Adults
1
* Environmental Toxicology Research Program and
Department of Pharmacology, The University of Mississippi, University, Mississippi 38677; and
Gulf Ecology Research Division, U. S. Environmental Protection Agency, Gulf Breeze, Florida 32561
Received October 19, 2001; accepted April 8, 2002
 |
ABSTRACT |
|---|
 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
17
-Ethinylestradiol (EE), a synthetic estrogen found in birth control pills, has been detected in the effluent of municipal wastewater treatment plants in several countries. Because EE was designed to be extremely potent at the estrogen receptor (ER), environmental exposure to low concentrations has the potential to disrupt the development of normal endocrine and reproductive function when exposure occurs during critical periods in development. Japanese medaka, Oryzias latipes, were used to evaluate the effect of exposure to EE during development on adult reproduction and endocrine function and the sensitivity of these animals to estrogen exposure as adults. To determine if the response to exogenous estrogen stimulation was diminished or sensitized, adults resulting from the developmental exposure groups were reexposed to EE at respectively higher concentrations. Hatchling exposure produced no changes in adult vitellogenin (VTG) content in the liver or circulating steroid concentrations, nor was reproduction affected. Reexposure of these adults inhibited reproduction, increased hepatic VTG and ER, and increased estrogen concentration measured in male plasma. Parental exposure produced permanent changes in hepatic content of ER and VTG in the adults resulting from exposure during gametogenesis and was related to a diminished response of males to subsequent estrogen exposure. The potential for this transgenerational exposure to decrease the responsiveness of males to EE is supported by comparing the concentration-response curves for hepatic VTG and ER in males exposed in ovo and as hatchlings. Our results indicate that the relationship between biomarkers and estrogen exposure will be altered by the timing and frequency of exposure. Key Words: teleost; vitellogenin; steroid; development; differentiation.
|
INTRODUCTION |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Greater awareness that synthetic chemicals may act as endogenous steroid hormones to disrupt critical developmental and reproductive processes (Coburn et al., 1996
) has led to a number of studies to identify environmental estrogens. Among those chemicals identified as environmental estrogens is the synthetic estrogen 17-ethinylestradiol (EE). EE is a potent estrogenic compound used extensively in oral contraceptive formulations, which are among some of the most widely prescribed pharmaceuticals (RxList, 1999). A recent investigation released by the UK Environment Agency reported that EE, endogenous estradiol (E2), and estrone significantly contributed to the estrogenic activity of select sewage effluents (Desbrow et al., 1998). Investigations conducted in several countries, including the United Kingdom, detected EE in sewage effluent at levels from 0.01 ng/l up to 7.0 ng/l (Desbrow et al., 1998; Ternes et al., 1999). Concerns regarding potential environmental exposure to pharmaceutical estrogens arise because of their greater potency relative to other well-studied environmental estrogens such as alkylphenol ethoxylates and organochlorine compounds (Nimrod and Benson, 1998).
Exposure to steroid hormones has a number of potential consequences for aquatic wildlife, the most severe of which is complete sex reversal. In Japanese medaka, the teleost fish Oryzias latipes, male phenotype is regulated by the presence of a Y chromosome. However, the sexual phenotype of many reptiles, birds, and fish, including medaka, can be altered by early life stage exposure to hormones. In some cases a complete reversal of sexual phenotype may occur (Hunter and Donaldson, 1983; Nimrod and Benson, 1998; Yamamoto, 1965). Steroid hormones have feedback mechanisms at the level of the pituitary and hypothalamus, but also act at target tissues to stimulate and maintain the reproductive tract and regulate gametogenesis in the gonads. Researchers have shown that development of intersex gonadal morphology is a sensitive biomarker to xenoestrogen exposure during development, with exposure beginning the day after hatching and continuing for 90 days (Metcalfe et al., 1999).
In oviparous animals, the liver is considered a target tissue for estrogens, as hepatocytes produce vitellogenin (VTG) in response to stimulation. (Ng and Idler, 1983). VTG is the glycophospholipoprotein precursor of egg yolk that provides nutrition for the developing embryo. As little as 2 ng/l EE induces VTG and inhibits testicular growth in adult male rainbow trout (Jobling et al., 1996). Therefore, it is quite possible that pharmaceutical products enter the aquatic environment in concentrations sufficient to elicit estrogenic responses. In fact, recent results from Metcalfe et al. (2001) have determined lowest-observed-effects concentrations resulting in developmental alterations of gonadal morphology to be below the reported environmental concentrations for some steroids, including EE. Ovarian tissue was observed in the testis of one male (of 33) exposed to 0.1 ng/l EE for 90 days.
Because of the role of natural estrogens in sexual differentiation (Crews, 1994), it is necessary to explore whether exposure to estrogenic chemicals during critical periods of differentiation can alter the permanence and severity of the consequences of exposure. Because steroid hormones are important in the development of the gonads, sexual differentiation, and gametogenesis, the question arises if a short exposure during development could permanently change the function of the reproductive system. The potential for an environmental estrogen to produce permanent changes in function or "imprint" the endocrine system would have serious implications for the impact of wastewater effluent on populations of wildlife. Early exposure (neonatally in mammals) has the potential to change the regulation of gene transcription, producing long-term and even epigenetic changes in response to excess hormonal signaling (McLachlan et al., 2001). These changes in gene regulation, or imprinting, have been implicated in the susceptibility to environmentally related diseases, including cancer (Jirtle et al., 2000). The earliest exposure of developing nonmammalian embryos to steroid hormones is the sequestering of maternal steroids in egg yolk prior to fertilization. Incorporation of maternal androgens and estrogens into eggs and the subsequent absorption of those steroids by the developing embryo have been well documented in birds (Schwabl, 1993, 1997). However, maternal steroids have also been documented in the eggs of turtles and fish (Bowden et al., 2001; Hwang et al., 1992).
The aim of this study is to compare the effects of EE exposure during the critical developmental periods of differentiation, beginning after 2 days posthatch (Nimrod and Benson, 1998), and parental or in ovo exposure on reproductive and endocrine function in Japanese medaka. Medaka were chosen as a model teleost because of their ability to reproduce consistently in the laboratory and their use as a model species for developmental and reproductive toxicity testing (Metcalfe et al., 1999). Developmental exposure either as hatchlings or in ovo allows us to determine persistent adult effects of developmental EE treatment. Some of these developmentally exposed adults were reexposed to EE to determine if early exposure permanently altered the type or magnitude of their estrogenic response. Adults were assessed for reproductive output and endocrine function, including circulating steroid concentrations, ex vivo steroidogenesis from the gonads, hepatic estrogen receptor (ER) content, and hepatic VTG. Physiological parameters can therefore be assessed for sensitivity to developmental exposure and as biomarkers of impaired reproduction.
|
MATERIALS AND METHODS |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Study Animals
Adult Japanese medaka, Oryzias latipes, have been maintained in culture at the University of Mississippi for approximately 7 years. Water quality parameters for Nanopure reconstituted balanced salt solution (BSS; Yamamoto, 1965
) were 82.5 mg/l CaCO3 hardness, 34.3 mg/l CaCO3 alkalinity, 6.637 pH, 23.5°C temperature, 8.83 mg/l dissolved oxygen, and 1.64 g/l salinity. Adult fish were fed twice daily with Tetramin flakes and brine shrimp in the morning and brine shrimp in the afternoon. Fry and hatchlings were fed only brine shrimp once or twice daily depending on their size. Cultured medaka were maintained at 26°C on a 15:9 light: dark cycle.
Hatchling exposure.
Previous unpublished data from a 96-h exposure of 2-day-old medaka fry to EE at concentrations ranging from 1 ng/l to 2000 ng/l indicate that the 96-h LC50 is greater than 2000 ng/l. Exposures in this study were performed at sublethal concentrations.
The day following hatching, medaka fry were placed in Nanopure BSS in a constantly lit incubator at 25°C. Beginning 2 days after hatching, medaka fry were exposed for 2 weeks to EE at nominal concentrations of 0, 0.2, 5, 500, and 2000 ng/l. Stock solutions of EE were made in 100% ethanol, and 25µl of these stocks were added to 500-ml jars containing 25 fry to achieve each nominal concentration. Water was completely renewed every 24 h. After the exposure, fish were transferred to 30-l tanks with untreated BSS and raised under normal laboratory conditions as described for the medaka culture. The fry were raised to adult size (approximately 3 months) and separated by sex.
Adult reproductive assessment and reexposure.
Adults resulting from the 0, 0.2, 5, 500, and 2000 ng/l exposure groups were paired, and their reproductive capacity was assessed (Fig. 1; Hatchling Exposure). Full-grown adults were placed in 800 ml exposure jars and set randomly in a water bath at 29°C maintained on a 15:9 h light:dark cycle. Prior to the reproductive assessment, individual adults were switched until all pairs were reproductively active. Approximately 2 h after the lights were turned on and the fish were fed, eggs were collected from each pair of adults. Eggs from each pair were maintained individually in 6-well plates and placed in an incubator at 25°C. Eggs were collected daily for 2 weeks and checked regularly for hatchlings, death, and algal and fungal growth for 30 days after collection. At the same time, adults resulting from the 0, 0.2, 5, 500, and 2000 ng/l groups were reexposed to EE at concentrations of 0, 0.8, 20, 2000, and 8000 ng/l, respectively, for 2 weeks and monitored as described (Fig. 1; Hatchling Exposure and Adult Reexposure). Therefore, a second control group, those pairs reexposed to the ethanol solvent as adults, is included in the analysis. Higher EE exposure concentrations were chosen in an attempt to determine if early exposure resulted in a desensitization to adult EE exposure. For each pair of adult animals, the total number of eggs, number of fertilized eggs, average egg size, date of first hatch, and number of hatchlings were recorded.
|
In ovo exposure.
Adult male-female pairs taken from the ongoing medaka culture were placed in 800 ml exposure jars and set randomly in a water bath at 29°C and maintained on a 15:9 h light:dark cycle. The pairs were acclimated for 3–4 days until egg production was consistent. If a specific pair did not reproduce prior to the start of the assessment, they were replaced before the exposure began. Stock solutions of EE were made in 100% ethanol, and 40 µl of these stocks were added to jars to achieve nominal concentrations of 0, 0.2, 5, 500, and 2000 ng/l. Every 24 h, when eggs were collected, the water was completely changed and reexposed to EE. The EE exposure did not result in any adult mortality.
Approximately 2 h after the lights were turned on and the fish were fed, eggs were collected from the individual pairs of adults. Eggs from each pair were maintained individually and checked regularly for hatchlings, death, and algal and fungal growth for 30 days after collection. The resulting offspring are considered to be exposed in ovo to EE to indicate that exposure occurred prior to fertilization. These animals exposed in ovo were raised to adult size (approximately 3 months) and separated by sex. The resulting in ovo animals were not sorted according to the length of the parental exposure to which they were subjected.
Adult reproductive assessment and reexposure.
Once full-grown, adult offspring were paired randomly and placed in 800-ml jars under the same conditions described above. Adults exposed in ovo appeared to be able to reproduce normally prior to the assessment. This transgenerational exposure contained eight treatment groups. Half of the adults resulting from the 0, 0.2, 5, 500, and 2000 ng/l exposure groups were paired, and their reproductive capacity was assessed (Fig. 1
; In Ovo Exposure). At the same time, the other half of the adults resulting from the 0, 0.2, 5, 500, and 2000 ng/l groups were reexposed as adults to EE at concentrations of 0, 0.8, 20, 2000, and 8000 ng/l, respectively (Fig. 1; In Ovo Exposure and Adult Reexposure). Higher EE concentrations were chosen in an attempt to determine if early exposure resulted in desensitization to adult EE exposure.
Reexposed adult pairs were monitored and exposed for 2 weeks with fresh stocks of EE made in ethanol. Mortality was observed among the pairs in the 8000 ng/l adult reexposure group. At the end of the 2-week exposure, three males and one female remained from the original four pairs of treated adults. Adults exposed in ovo were placed in the same configuration, and reproduction was monitored for a 2-week period without exposure. For each pair of adult animals, the total number of eggs, number of fertilized eggs, average egg size, date of first hatch, and number of hatchlings were recorded.
Physiological Parameters
After reproductive assessment was complete, tissue samples were collected to perform several assays of endocrine function. Each fish was anesthetized with MS-222 (3-aminobenzoic acid ethyl ester, methanesulfonate salt, Sigma Chemical, St. Louis, MO), weighed, and measured, and blood was collected by cutting the isthmus above the heart and placing a 5-µl Drummond Scientific heparinized, disposable micropipet (Fisher Scientific, St. Louis, MO) in the flow of arterial blood from the heart. Whole blood was pooled from two individuals of the same sex and transferred to a microcentrifuge tube that contained 2 µl of a 6.5 mg/ml solution of sodium heparin salt in water. Pooled blood samples were spun in a 4°C centrifuge (6000–10,000 rpm; 10 min) to separate the plasma. The plasma was measured and transferred into microcentrifuge tubes containing PMSF (99% purity, phenylmethylsulfonyl fluoride, Sigma) and stored in a –80°C until analyzed. The livers were removed from the same two individuals, pooled, and stored in microcentrifuge tubes containing PMSF. Gonads were removed and put in a Falcon Pro-Bind 96-well, flat-bottomed plate containing 100 µl Media 199 (with Hanks’ salts, L-glutamine, and without sodium bicarbonate; Gibco Life Technologies, Grand Island, NY) and 2 µl/gonad of a 0.01mg/ml concentration of 25-hydroxycholesterol. The media was removed and stored at –80°C freezer until analyzed. Whole-liver homogenates were prepared by homogenizing pooled samples and spinning out cellular debris in a 4°C centrifuge (10,000–13,000 rpm; 30 min). The supernatant was removed and protein content was measured. Protein concentrations were determined in 96-well plates using Bio-Rad Protein Assay protein dye and standards of 0.5, 0.25, 0.125, and 0.0625 mg/ml bovine serum albumin (BSA; Sigma, cold alcohol precipitated, frac V). Concentrations were used to assay VTG and ER in samples with standard protein content.
Whole-liver homogenates were analyzed for VTG and ER content by Western blot. Determination of VTG in liver samples was used as an endpoint to conserve the small volumes of collected plasma for steroid analysis. Further tissue separation to produce microsomes was not done with medaka liver samples because of restricted tissue volume and because cytosolic separation of pooled liver samples proved to have similar ER banding as whole-liver samples. VTG- and ER-positive tissue samples were made by exposing mature male and female medaka to 50 µg/l 17ß-estradiol for 1 week with 24 h, 100% static renewal. Livers from all exposed animals were pooled, protein concentration was determined, and the positive tissue was aliquoted into samples to be used as a positive control on each gel.
VTG was determined using a monoclonal anti-VTG antibody developed by Heppell et al. (1995) against striped bass VTG and purchased from Cayman Chemical (Ann Arbor, MI). This anti-VTG antibody identifies a 170-kDa band in striped bass (Thompson, 2000) and two distinct bands in medaka plasma samples (Gronen et al., 1999). A hepatic cytosolic assay for VTG in medaka with the use of this antibody was demonstrated by Nimrod and Benson (1997); this measurement was determined to show a clear dose response to aqueous E2 exposure. In hepatic samples, VTG was identified as a primary band of 200 kDa and a second fainter 120-kDa band in several samples, and was quantified as the integrated optical density contained in both bands (Fig. 2A). Some faint bands, presumably breakdown products, were visible in liver samples. When these were present, the optical density of these bands was included in the integrated measurement.
|
The primary antibody for ER detection was made in mouse against the DNA-binding domain of the ER in humans (catalogue number MA1-310, Affinity Bioreagents, Inc., Stillwater, MN). ER was identified as a single, approximately 65-kDa band for medaka, although two additional bands, one slightly smaller and one at approximately 130 kDa, were often visible. All bands present were included in the measurement of integrated optical density. A representative ER gel is shown in Figure 2B
. After the nitrocellulose image of each Western blot was developed, it was scanned and analyzed using Scion Image software (Scion, Frederick, MD). Integrated optical density (IOD) of each lane was recorded and compared with the optical density of the same positive tissue sample that was run on each gel. Each sample had a recorded (nonzero) IOD. Samples with no bands had an IOD equivalent to that of background and resulted in a low relative IOD to the positive tissue sample. However, normalizing across gels for the same positive control ensured that each sample had a numerical value for ER and VTG content. Circulating steroid concentrations were determined by extracting the plasma steroids from pooled blood samples and using a steroid enzyme immunosorbant assay (EIA) developed by Munro and Lasley (1998). The plasma volume was measured with a Hamilton syringe, transferred to 10-ml test tubes, and extracted three times with ethyl acetate. For each extraction, 100 µl ethyl acetate was added to the sample, mixed for 5 s, and spun at 4000 rpm for 5–10 min. After each centrifugation, the organic layer was transferred into a new test tube. The combined ethyl acetate extracts were allowed to dry overnight or placed on a nitrogen blower until completely dry. Standards of both E2 and T in ethyl acetate at levels of 5, 10, 25, 50, 100, and 150 pg, and a control of ethyl acetate were used as a standard curve. A 96-well, flat-bottomed plate was coated 24 h before use with either 1:10,000 anti-E2 antibody or 1:20,000 anti-T antibody. Samples and standards were reconstituted in phosphate-buffered saline with BSA and then added to the plate. A standard dilution of steroid (either E2 or T) conjugated to horseradish peroxidase (HRP) was then added to each well. Two hours after incubation of sample or standard and HRP conjugate, the plate was washed and an HRP substrate added. Plates were then read at 405 nm (570 nm reference).
As an assay of potential changes in gonadal physiology in response to steroid exposure, gonadal steroidogenesis was determined by the ex vivo release of E2 and T in the presence of 25-hydroxycholesterol, similar to methods described by Srivastava and Van der Kraak (1994). Testis and ovaries were removed intact from fish and put in a Falcon Pro-Bind 96-well, flat-bottomed plate containing 2 µl/gonad of a 0.01 mg/ml concentration of 25-hydroxycholesterol and 100 µl of Media 199 (GIBCO Life Technologies, with Hanks’ salts, L-glutamine, and without sodium bicarbonate) supplemented with 25 mM Hepes, 4.0 mM sodium bicarbonate, 0.01% streptomycin sulfate, and 0.1% BSA, as described by Van der Kraak, et al. (1992).
An initial trial determined the time-dependent and concentration-dependent nature of E2 production from 25-hydroxycholesterol-stimulated ovaries and testes compared with basal levels (Thompson, 2000). Gonads were incubated in 100 µl Media 199 with 2 µl of 0, 4.12, or 41.2 µM 25-hydroxycholesterol in ethanol for 24 or 48 h. A subset of gonads were prerinsed in media for 2 h prior to the addition of 25-hydroxycholesterol to ensure that circulating steroid concentrations did not alter the measurement of steroid release from the gonads. 25-Hydroxycholesterol increased ex vivo E2 release in a time- and concentration-dependent manner. Prerinsing had no effect on E2 production (Thompson, 2000). Optimal conditions for both ovaries and testis were determined to be a 48-h incubation with 4.12 mM 25-hydroxycholesterol. A second trial tested the variability of ex vivo E2 production in 25-hydroxycholesterol-stimulated gonads compared with basal levels. Replicates of four ovaries and testes were treated with 2 µl ethanol or with 4.12 mM 25-hydroxycholesterol. Addition of 25-hydroxycholesterol increased E2 production 4-fold in testes and 2-fold in ovaries (Thompson, 2000).
Ex vivo steroidogenesis was determined in EE experimental animals with a 48-h incubation at room temperature with 4.12 mM 25-hydroxycholesterol. Following the incubation, steroid amounts released by gonads were quantified by extraction of the media with ethyl acetate and determining E2 and T content using the described steroid EIA.
Statistics
Comparisons across groups within an experiment were made with a one-way ANOVA followed by post hoc pairwise comparisons using Fisher’s protected least significant difference (PLSD) tests. Percent of eggs that were fertilized and percent of fertilized eggs that hatched were analyzed as nonparametric variables using Kruskal-Wallis comparisons. A 2 x 2 contingency table was used to detect significant changes in sex ratio. Concentration-response relationships were plotted using GraphPad Prism (GaphPad Software, San Diego, CA) to find the best fit of EE exposure concentrations and the IOD from Western blot analysis relative to positive tissue samples.
|
RESULTS |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Hatchling Exposure
Sex ratio of developmentally exposed adults.
The sex ratio of groups of adults exposed for 2 weeks to EE as hatchlings contained more females at every treatment level, although the ratio was not significantly different from controls. The solvent-treated control group contained 17 females and 15 males. For EE treatments, the number of females and males were as follows: 0.2 ng/l, 19 females, 15 males; 5.0 ng/l, 17 females, 15 males; 500 ng/l, 23 females, 9 males; and 2000 ng/l, 17 females, 9 males. A 2 x 2 contingency table indicated that the female bias in sex ratio for the 2000 ng/l treated medaka was not significantly different from controls (
2c = 0.158, df = 1, p > 0.05). However, the trend of a female bias in the resulting sex ratio with higher EE concentrations indicates that some of the resulting "females" may be sex-reversed males.
Reproduction.
Adult Japanese medaka exposed as hatchlings did not exhibit any changes in reproduction (Table 1). However, reexposure of adult animals to EE did alter reproductive endpoints relative to the solvent reexposed control group. The number of eggs produced during a 2-week period did not change for animals exposed as hatchlings, but did decrease with the 2000 ng/l and 8000 ng/l reexposure (ANOVA, F = 18.99, p < 0.0001; Fisher’s PLSD post hoc pairwise tests, p = < 0.001; Fig. 3). The percent of all eggs that were fertilized was unchanged with either exposure of hatchlings or adult reexposure (Kruskal-Wallis test, H = 10.88, p = 0.28). The size of eggs produced by females increased with reexposure to 2000 ng/l EE (ANOVA, F = 2.56, p = 0.02; Fisher’s PLSD post hoc pairwise tests, p = 0.02). However, the size of eggs did not translate to a change in the percentage of fertilized eggs that hatched within a month of being laid (Kruskal-Wallis test, H = 11.32, p = 0.25).
|
|
Physiological Parameters
The ex vivo steroidogenesis of the gonads of treated animals was not changed either by EE exposure at the hatchling stage or as adults. Production of E2 from ovaries was unchanged (Table 2
; ANOVA, F = 0.37, p = 0.94), and production from testis tended to decrease with increasing exposure, although the change was not significant (ANOVA, F = 1.11, p = 0.35). Production of testosterone was unchanged from either testis or ovaries (ANOVA, F < 0.48, p > 0.88).
|
The response of circulating steroid concentrations differed between females and males (Table 2
). Among females, there was no change in circulating E2 concentrations (ANOVA, F = 1.48, p = 0.19). The testosterone concentration in plasma tended to increase with reexposure, although none of the increases were significantly different from controls (ANOVA, F = 2.59, p = 0.02; Fisher’s PLSD tests, p > 0.05). Among males, there was a significant increase in plasma E2 concentration with adult reexposure at 2000 ng/l and 8000 ng/l (ANOVA, F = 3.09, p = 0.02; Fisher’s PLSD tests, p > 0.03) and a tendency to increase with 500 ng/l of EE at the hatchling stage (Fisher’s PLSD tests, p = 0.07; Fig. 4). There was no change in circulating testosterone in male plasma. However, there was a tendency for testosterone to decrease with both 5 ng/l exposures at the hatchling stage and 20 ng/l reexposure as adults (ANOVA, F = 2.01, p = 0.09).
|
Hepatic ER concentrations were measured with antihuman ER antibody. The bands found were at approximately 65 kDa, the expected size of the ER from work in other fishes (Pakdel et al., 1994
; Todo et al., 1996) and the ER subtype from medaka (Tchoudakova et al., 1999), and a band of approximately 130 kDa, twice the expected size. The larger band may represent nonspecific binding of this antibody to another larger nuclear receptor or potentially the detection of ER dimers. Only one ER sequence has been reported for Japanese medaka (ER; Okada et al., 1996). Given that the antihuman ER antibody recognizes the DNA-binding domain, which is expected to be highly conserved, another possibility is that the bands detected in this study may represent multiple ER forms.
The response of hepatic ER content was different for the hatchling exposure and adult reexposure. Among males, ER content decreased significantly with hatchling exposure at 500 ng/l and 2000 ng/l (ANOVA, F = 5.01, p = 0.0007; Fisher’s PLSD tests, p < 0.01) and increased significantly with adult reexposure at 8000 ng/l (Fisher’s PLSD tests, p = 0.0001; Fig. 5A). Females exhibited the same trend, with hatchling exposure decreasing ER content and reexposure increasing ER content, although the changes were not significant (ANOVA, F = 1.72, p = 0.12; Fig. 5B).
|
VTG content in the liver of males and females exposed to EE changed only with adult reexposure. Among females, reexposure to 2000 ng/l and 8000 ng/l increased hepatic VTG (ANOVA, F = 3.23, p = 0.005; Fisher’s PLSD tests, p < 0.02; Fig. 6
). Among males, 2000 ng/l and 8000 ng/l reexposure as adults significantly increased hepatic VTG (ANOVA, F = 13.04, p < 0.0001; Fisher’s PLSD tests, p < 0.03; Fig. 6).
|
In Ovo Exposure
Adult sex ratio.
In ovo exposure did not result in a significant change in the sex ratio of the resulting offspring relative to the control group. Thus, transgenerational exposure did not appear to cause sex change in developing embryos. The numbers of females and males from each treatment group were as follows: control, 105 female, 76 male; 0.2 ng/l, 99 female, 87 male; 5.0 ng/l, 92 female, 87 male; 500 ng/l, 36 female, 33 male; and 2000 ng/l, 8 female, 15 male.
Reproduction.
Several measures of reproductive function were impaired by adult reexposure to EE at higher concentrations, but no difference was detected between animals from EE in ovo exposures and the solvent control group (Table 1). Egg production during the 2-week reexposure was lower in the 2000 ng/l and 8000 ng/l EE reexposure group relative to reexposed controls, but was unchanged within the in ovo exposure groups (ANOVA, F = 15.45, p < 0.0001; Fisher’s PLSD tests, p < 0.0001; Fig. 7). The average size of eggs produced by females also increased with reexposure to higher concentrations of EE (ANOVA, F = 12.40, p < 0.0001; Fisher’s PLSD tests, p < 0.004; Fig. 8). The number of hatchlings produced by each pair paralleled the number of eggs produced, except for those treatment groups that exhibited changes in rates of fertilization and hatching success of fertilized eggs. There was no effect on the rate of fertilization within the EE exposure treatments either in ovo or with reexposure groups (Kruskal-Wallis test, H = 6.71, p > 0.67). The highest concentration of reexposure to EE showed a decrease the percent of fertilized eggs that hatched within 30 days of fertilization (Kruskal-Wallis test, H = 21.7, p = 0.01); this same trend was seen in the animals exposed in ovo from the 2000 ng/l treatment group (Fig. 9). Hatched offspring appeared normal, but viability was not assessed.
|
|
|
Physiological Endpoints
Circulating testosterone concentration in males was stimulated by adult reexposure to EE (Table 3
). However, other measures of plasma steroids were unaffected by EE treatment. Among females, circulating E2 (ANOVA, F = 0.599, p = 0.77) and testosterone (ANOVA, F = 0.571, p = 0.80) were not altered by EE treatment either in ovo or with adult reexposure. Among males, E2 in plasma tended to increase with 2000 ng/l in ovo exposure and 8000 ng/l adult reexposure relative to their respective controls, although the difference was not significant (ANOVA, F = 1.16, p = 0.35). However, circulating testosterone in males with 8000 ng/l of reexposure, although highly variable, did significantly increase relative to all other reexposure groups, including controls (ANOVA, F = 2.07, p = 0.05; Fisher’s PLSD tests, p < 0.008; Fig. 10).
|
|
Ex vivo steroidogenesis from ovaries incubated with 0.01 mg/ml 25-hydroxycholesterol did not change within EE treatments in ovo or with reexposure of adult (Table 3
). However, there was a trend of decreased E2 production from ovaries of females reexposed as adults (ANOVA, F = 1.69, p = 0.13). Production of testosterone from ovaries tended to decrease with both in ovo exposure to EE at both 500 ng/l and 2000 ng/l and with reexposure to 2000 ng/l and 8000 ng/l relative to their respective control groups (ANOVA, F = 1.71, p = 0.13). Ex vivo release of E2 from testis was not altered by EE treatment at any stage (ANOVA, F = 1.16, p = 0.34). However, production of testosterone from testis did appear to be inhibited by reexposure to 2000 ng/l and 8000 ng/l, although the difference was not significant (ANOVA, F = 1.28, p = 0.27).
The relationship between EE exposure and hepatic ER content differed between males and females. ER in liver did not differ among females regardless of treatment (ANOVA, F = 1.27, p = 0.28). Among liver samples from males, there was a significant increase in ER in males exposed in ovo to 2000 ng/l, and those reexposed to 8000 ng/l (ANOVA, F = 2.10, p = 0.05; Fisher’s PLSD tests, p < 0.05; Fig. 11).
|
VTG content of the liver also increased with certain EE exposures. Specifically, VTG content in the liver was elevated with 0.2 ng/l in ovo exposure of females and 2000 ng/l reexposure of adult females (ANOVA, F = 7.08, p < 0.0001; Fisher’s PLSD tests, p < 0.003; Fig. 12
). Among males, VTG content was elevated significantly only at 8000 ng/l reexposure (ANOVA, F = 3.46, p = 0.003; Fisher’s PLSD tests, p < 0.0001; Fig. 12).
|
|
DISCUSSION |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Hatchling Exposure
Exposure of adult Japanese medaka to EE has been reported to have a number of consequences, including sex reversal, decreased female fecundity, expression of aromatase in the gonads (Scholz and Gutzeit, 2000
), and the development of gonadal alteration or testis-ova in males (Metcalfe et al., 2001). Adult exposure to 14 days of EE at nominal concentrations of 0, 0.2, 5.0, 500, and 2000 ng/l resulted in both physiological and reproductive changes (Thompson, 2000). EE exposure induced plasma VTG and hepatic ER protein concentrations in males and females at 500 ng/l and 2000 ng/l. Plasma E2 concentration was elevated for males and females at exposure concentration of 5 ng/l and above, although no change in circulating testosterone was reported. Ex vivo synthesis of E2 was elevated in both ovaries and testis when animals were exposed to 0.2 ng/l or 500 ng/l of EE. Ex vivo synthesis of testosterone was reduced in both ovaries and testis at concentrations of 500 ng/l.
In our study, adults exposed as hatchlings to EE showed no reproductive impairment. However, reexposure of adults to EE led to a significant decrease in egg production at higher concentrations. This finding parallels the results of adult-only exposure to EE in the same conditions (Thompson, 2000). Treatment of adults with EE at 5 ng/l for 2 weeks did not result in decreased fecundity, whereas 500 ng/l and 2000 ng/l concentrations were associated with decreased egg production. Scholz and Guteit (2000) have observed significant decreases in the number of spawning adult females (and egg production per female) after developmental exposure to EE at 10 ng/l for 2 months starting immediately after hatching. Therefore, whereas exposure of medaka to higher EE concentrations for 2 weeks produced no permanent effect, exposure to lower concentrations of EE for 2 months produced persistent changes in fecundity. These results suggest that either the critical window for gonadal development is longer than the 2-week exposure period, or that the duration, more than concentration of developmental exposure, might influence female fecundity.
Developmental exposure of medaka to estrogens including EE is known to produce testis-ova in some males and complete sex reversal in others (Metcalfe et al., 2001; Nimrod and Benson, 1998; Scholz and Gutzeit, 2000). The duration of exposure in these studies ranges from 1 to 3 months posthatch. Although trends in sex ratio may indicate some male to female sex change with this 2-week exposure, no significant deviation from 50% female population was detected. In addition, any reversal of sex from male to female must have resulted in functional female medaka, as no decline in reproduction was observed after exposure as hatchlings. In mammals, EE treatment results in a decrease in sperm motility associated with declining testosterone (Kaneto et al., 1999). The results presented here provide no evidence that a 2-week exposure produces permanent sex change in males. Not only was the ratio of males to females unaffected by EE in this study, the fecundity of females (egg production) and the fertility of males (percent of eggs fertilized) was similar in all animals exposed as hatchlings. At least two potential explanations exist for the absence of noticeable sex change in hatchlings exposed to EE. First, the 2-week exposure beginning at 2 days posthatch used in this study may not have completely spanned the critical range for gonadal alteration (testis-ova). Gonadal development may have a period of growth early, but determination of testis-ova tissue may span a longer time frame. Thus early developmental steroid exposure would be critical to cause sex change, but extended exposure is necessary to see phenotypic changes. Alternatively, early steroid exposure may produce gonadal sex change, and removal of exogenous steroids allows reversal of ovarian tissue back to testis structure. The 2-month growth period after exposure may have been adequate time for testis tissue to recover to normal function prior to reproductive testing.
In response to EE stimulation, steroid parameters were relatively unchanged in these assays. Neither circulating steroid concentrations or ex vivo steroid release from the gonad was altered in females exposed to EE as hatchlings or as adults. In males, plasma E2 concentrations were increased with adult reexposure to higher concentrations of EE. Adult exposure to EE as described by Thompson (2000) showed a resulting increase in circulating E2 in both sexes and associated changes in steroidogenesis. Other researchers have found that E2 treatment of medaka with 2.5 µg/l for 10 days altered the activity of the aromatase enzyme responsible for conversion of testosterone to E2 in the central nervous system (Melo and Ramsdell, 2001). In comparison, few changes were detected in the steroid measurements of adults exposed as hatchlings.
Developmentally treated adult males had elevated hepatic VTG and ER with EE treatment at micrograms per liter concentrations. Other research has demonstrated an induction of hepatic VTG expression and correlative induction in plasma VTG concentrations in adult male sheepshead minnows in response to EE treatments (Bowman et al., 2000; Denslow et al., 2001). In medaka, a single adult treatment with EE produced elevated VTG and ER in both sexes at 500 ng/l (Thompson, 2000). Therefore, although female VTG induction was not observed in this situation, there is no evidence that exposure as hatchlings has altered the male vitellogenic response to EE. Surprisingly, hepatic ER content in male and female controls exposed to ethanol as hatchlings tended to be elevated, although nonsignificantly, above controls treated again with ethanol as adults. The difference between these two control groups is reexposure to the solvent (50 µl/l ethanol) during the 2-week reproductive assessment of adult animals. Because the controls exposed as hatchlings have hepatic ER contents that are greater than the positive tissue (liver samples from adults exposed for 7 days to 50 µg/l E2), it is proposed that these samples have elevated ER content relative to reexposed controls. However, we cannot rule out the possibility that ER content is reduced with adult reexposure to ethanol. Though the difference between these controls does not reach statistical significance, it does suggest a role for ethanol in ER regulation.
In Ovo Exposure
Although hatchling exposure to estrogens has been reported to produce complete male to female sex reversal in medaka (Metcalfe et al., 2001; Nimrod and Benson, 1998; Scholz and Gutzeit, 2000), in ovo exposure to EE did not result in a female-biased sex ratio in full-grown animals. Furthermore, reproduction in female adults exposed transgenerationally to EE did not result in impairment of reproductive function in terms of fecundity or egg size. However, at the highest concentration of in ovo exposure, there was a trend for fewer fertilized eggs to produce viable hatchlings, and this pattern reached statistical significance with repeated EE exposure to adult animals.
Reexposure as adults of these transgenerationally treated animals inhibited reproduction at concentrations of 2000 ng/l EE and decreased the percentage of fertilized eggs that produced viable embryos at even higher concentrations. The effective concentration of EE in this reexposure is comparable to the concentrations that caused reproductive inhibition in untreated adults in a related study (Thompson, 2000). Hatchling success of fertilized eggs produced by adults exposed to EE in that study was impaired, relative to controls, at concentrations of 500 ng/l but not 2000 ng/l. A decrease in egg number and hatchling production has been demonstrated for aquatic E2 exposure and other xenoestrogens using medaka by Shioda and Wakabayashi (2000). Hence, this 2-week in ovo exposure is not linked to any permanent change in reproduction or any obvious change in the threshold or type of reproductive impairment caused by adult exposure.
The measured parameters for steroid hormones were unchanged for most of the measurements in this exposure. EE treatment either in ovo or as adults had no effect on the steroidogenesis of the ovaries or the testes ex vivo. Circulating concentrations of plasma hormones were unaffected for E2. However, male plasma testosterone concentrations were elevated with the highest reexposure of adult animals. With simple adult exposure, E2 released from ex vivo gonadal tissue as well as plasma concentrations increase with exposure to relatively low levels of EE (5 ng/l; Thompson, 2000). In Thompson’s findings, testosterone released from testes showed an interesting pattern of increasing at low exposure concentrations (5 ng/l) and decreasing at higher concentrations (500 ng/l). Relative to adult exposure, transgenerational exposure produces a diminished response or a buffering in steroidogenesis and circulating steroid concentrations after exposure to estrogens.
Hepatic ER and VTG significantly change with transgenerational exposure to EE at specific concentrations. Among females, low concentrations of EE resulted in elevated VTG in adult animals, and the concentration of the ER in the liver tended to decrease with in ovo exposure. Among males, the relationship between in ovo exposure and hepatic ER content was reversed. In ovo exposure to 2000 ng/l EE caused an induction of ER in the liver of the male offspring. Changes in the liver as a result of in ovo exposure are likely to be permanent or imprinted responses to early treatment with estrogens. Although these changes are not correlated with reproductive impairment, changes in VTG or ER in the liver may be related to the liver function and have the potential to change the relationship between exposure and biomarkers.
The reexposure of transgenerationally treated males resulted in an induction of hepatic VTG and ER at the highest EE concentrations. Exposure to 8000 ng/l was necessary to produce significant VTG induction (2000 ng/l did not) for adult males exposed in ovo. Compared with the results from Thompson (2000) for a single adult exposure, in which VTG and ER induction occurred at 500 ng/l, the response of males exposed in ovo was much less sensitive to EE treatment. A comparison across the two different exposure regimes in this study of the response to EE treatment, that of males exposed in ovo to males exposed as hatchlings, suggests that in ovo exposure diminishes the magnitude of the estrogenic response of adults reexposed to EE. The concentration-response curve plotted for hepatic ER (Fig. 13) and hepatic VTG (Fig. 14) indicates that the induction (relative to the same positive tissue samples run on each gel) is greater in response to EE reexposure for males treated with EE as hatchlings than for males exposed in ovo. One hypothesis for the diminished response of animals treated in ovo is that the presence of sequestered steroids in egg yolk changes the threshold for triggering an estrogenic response.
|
|
Conclusions
Partial life cycle and complete life cycles assessments of the effects of EE on development and reproduction have been completed in sheepshead minnow (Zillioux et al., 2001
) and fathead minnows (Länge et al., 2001). Partial life-cycle exposure of sheepshead minnow fry (Zillioux et al., 2001) for 59 days until subadult stages resulted in gonadal abnormalities in males from 200 ng/l treatment groups as well as a decrease in fecundity of females and a drop in hatching success at that exposure level. Full life-cycle assessment in fathead minnows (Länge et al., 2001) involved a 305-day exposure, with treatment beginning with newly fertilized embryos. With full life-cycle exposure, 100% of animals exposed to 4 ng/l EE were found to be female by histological examination of the gonads. The offspring of adults exposed at concentration lower than 4 ng/l did not exhibit a female-biased sex ratio, although the offspring in these exposure groups were smaller than control fry. Two conclusions from these studies support the results presented here. First, although exposure to EE impacts reproduction for exposed animals, there is no indication that parental exposure will result in changes in reproductive traits in the next generation. Second, the more prolonged of these exposures was associated with a lower concentration of EE for observed effects (200 ng/l in a 59-day exposure and 4 ng/l in a 305-day exposure). The duration of exposure may be inversely associated with the sensitivity of the response. Adult medaka exhibit change in reproductive and physiological parameters after 14 days of exposure to 500 ng/l of EE (Thompson, 2000). However, 90 days of exposure to EE at concentrations as low as 0.1 ng/l can result in the formation of testis-ova (Metcalfe et al., 2001). Therefore, species sensitivity to EE can be compared only across homologous exposure scenarios. Parental treatment with xenoestrogens has the potential to produce permanent changes in the endocrine function of the offspring they produce. In this study, we have found that in ovo treatment with EE does not alter the reproductive capacity of Japanese medaka or the changes in physiology associated with adult estrogen exposure and reproduction described by Thompson (2000). However, parental treatment did change the relationship between adult exposure and physiological endpoints. Specifically, adult exposure was less likely to change steroid measurements, including steroidogenesis and circulating steroid concentrations. Furthermore, parental exposure produced permanent changes in hepatic content of ER and VTG, as evidenced by changes in hepatic ER content in males from parents exposed to 2000 ng/l EE, and hepatic VTG in females from parents exposed to 2 ng/l of EE. Parental exposure was also related to a diminishing response of males to estrogen exposure relative to the response of males treated as hatchlings and the reported threshold values for adult medaka. The potential for transgenerational exposure to decrease the responsiveness of males to EE is supported by comparing the concentration-response curves for hepatic VTG and ER in males exposed in ovo and as hatchlings. Our results indicate that the relationship between biomarkers and estrogen exposure will be further complicated by the timing and frequency of exposure.
These findings suggest 2 weeks of EE exposure developmentally is unlikely to result in reproductive impairment of one specific age class of fish, although high concentrations of EE may alter adult reproductive function. Developmental exposure may result in alteration or impairment of gonadal function only if exposure continues through gonadal differentiation for an extended period. Measurements of VTG seem to be reliable indicators of current estrogenic loads and may correlate with measures of fecundity in medaka. Importantly, the relationship between certain biomarkers, specifically steroid measurements, and impairment may be diminished with developmental exposure.
|
ACKNOWLEDGMENTS |
|---|
We would like to thank Alison Nimrod for help with the experimental design, Susan Thompson for her work in developing and testing several assays, and K. Erica March for her help with the medaka culture. We also wish to acknowledge the helpful comments of the initial reviewers of this manuscript. This research was supported in part by the Environmental and Community Health Research program at the University of Mississippi and by the U.S. Environmental Protection Agency (R82-7098-010).
|
NOTES |
|---|
1 To whom correspondence should be sent at present address: Department of Biology, West Virginia University, Morgantown, WV 26506-6057. Fax: (304) 293-6363. E-mail: cmforan{at}mail.wvu.edu.
This research was supported in part by the U.S. Environmental Protection Agency. It has not been subjected to the Agency’s peer and policy review and therefore does not necessarily reflect the views of the Agency. No official endorsement should be inferred.
|
REFERENCES |
|---|
|
TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONREFERENCES |
|---|
Bowden, R. M., Ewert, M. A., Lipar, J. L., and Nelson, C. E. (2001). Concentrations of steroid hormones in layers and biopsies of chelonian egg yolks. Gen. Comp. Endocrinol. 121, 95–103.[Web of Science][Medline]
Bowman, C. J., Kroll, K. J., Hemmer, M. J., Folmar, L. C., and Denslow, N. D. (2000). Estrogen-induced vitellogenin mRNA and protein in sheepshead minnow (Cyprinodon variegates). Gen. Comp. Endocrinol. 120, 300–313.[Web of Science][Medline]
Coburn, T., Dumanoski, D., and Myers, J. P. (1996). Our Stolen Future. Penguin Books, New York.
Crews, D. (1994). Animal sexuality. Sci. Am. 270, 109–114.[Medline]
Denslow, N. D., Bowman, C. J., Ferguson, R. J., Lee, H. S., Hemmer, M. J., and Folmar, L. C. (2001). Induction of gene expression in sheepshead minnows (Cyprinodon variegates) treated with 17ß-estradiol, diethylstilbestrol, or ethinylestradiol: The use of mRNA fingerprints as an indicator of gene regulation. Gen. Comp. Endocrinol. 121, 250–260.[Web of Science][Medline]
Desbrow, C., Routledge, E. J., Brighty, G. C., Sumpter, J. P., and Waldock, M. (1998). Identification of estrogenic chemicals in STW effluent. I. Chemical fractionation and in vitro biological screening. Environ. Sci. Technol. 32, 1549–1558.
Gronen, S., Denslow, N., Manning, S. Barnes, S., Barnes, D. and Brouwer, M. (1999). vitellogenin levels and reproductive impairment of male Japanese medaka (oryzias latipes) exposed to 4-tert-octylphenol. Environ. Health Perspect. 107, 385–390.[Web of Science][Medline]
Heppell, S. A., Denslow, N. D., Folmar, L. C., and Sullivan, C. V. (1995). Universal assay of vitellogenin as a biomarker for environmental estrogens. Environ. Health Perspect. 103(Suppl. 7), 9–15.
Hunter, G. A., and Donaldson, E. M. (1983). Hormonal sex control and its application to fish culture. In Reproduction, Part B: Behavior and Fertility Control IX (W. S. Hoar, D. J. Randall, and E. M. Donaldson, Eds.), pp. 223–303. Academic Press, Orlando, FL.
Hwang, P. P., Wu, S. M., Lin, J. H., and Wu, L. S. (1992). Cortisol content of eggs and larvae of teleosts. Gen. Comp. Endocrinol. 86, 189–196.[Web of Science][Medline]
Jirtle, R. L., Sander, M., and Barrett, J. C. (2000). Genomic imprinting and environmental disease susceptibility. Environ. Health Perspect. 108, 271–278.[Medline]
Jobling, S., Sheaham, D., Osborne, J. A., Matthiessen, P., and Sumptor, J. P. (1996). Inhibition of testicular growth in rainbow trout (Oncorhynchus mykiss) exposed to alkylphenolic chemicals. Environ. Toxicol. Chem. 15, 194–202.
Kaneto, M., Kanamori, S., Hishikawa, A., and Kishi, K. (1999). Epididymal sperm motion as a parameter of male reproductive toxicity: Sperm motion, fertility, and histopathology in ethinylestradiol-treated rats. Reprod. Toxicol. 13, 279–289.[Web of Science][Medline]
Länge, R., Hutchinson, T. H., Croudace, C. P., Siegmund, F., Schweinfurth, H., Hampe, P., Panter, G. H., and Sumpter, J. P. (2001). Effects of the synthetic estrogen 17-ethinylestradiol on the life-cycle of the fathead minnow (Pimephales promelas). Environ. Toxicol. Chem. 20, 1216–1227.[Web of Science][Medline]
McLachlan, J. A., Burow, M., Chaing, T. C., and Li, S. F. (2001). Gene imprinting in developmental toxicology: A possible interface between physiology and pathology. Toxicol. Lett. 120, 161–164.[Web of Science][Medline]
Melo, A. C., and Ramsdell, J. S. (2001). Sexual dimorphism of brain aromatase activity in medaka: Induction of a female phenotype by estradiol. Environ. Health Perspect. 109, 257–264.[Web of Science][Medline]
Metcalfe, C. D., Gray, M. A., and Kiparissis, Y. (1999). The Japanese medaka (Oryzias latipes): An in vivo model for assessing the impacts of aquatic contaminants on the reproductive success of fish. In Impact Assessment of Hazardous Aquatic Contaminants (S. S. Rao, Ed.), pp. 29–52. Lewis Publishers, Boca Raton, FL.
Metcalfe, C. D., Metcalfe, T. L., Kiparissis, Y., Koenig, B. G., Khan, C., Hughes, R. J., Croley, T. R., March, R. E., and Potter, T. (2001). Estrogenic potency of chemicals detected in sewage treatment plant effluents as determined by in vivo assays with Japanese medaka (Oryzias latipes). Environ. Toxicol. Chem. 20, 297–308.[Web of Science][Medline]
Munro, C. J., and Lasley, B. L. (1988). Non-radiometric methods for immunoassay of steroid hormones. Prog. Clin. Biol. Res. 285, 289–329.[Medline]
Ng, T. B., and Idler, I. R. (1983). Yolk formation and differentiation in teleost fishes. In Fish Physiology IX (W. S. Hoar, D. J. Randall, and E. M. Donaldson, Eds.), pp. 373–404. Academic Press, Orlando, FL.
Nimrod, A. C., and Benson, W. H. (1997). Assessment of estrogenic activity in fish. In Chemically Induced Alterations in Functional Development and Reproduction of Fishes (R. M. Rolland, M. Gilbertson, and R. E. Peterson, Eds.), pp. 87–100. SETAC Publishing, Pensacola, FL.
Nimrod, A. C., and Benson, W. H. (1998). Reproduction and development of Japanese medaka following early life stage exposure to xenoestrogens. Aquat. Toxicol. 44, 141–156.
Okada, H., Kawahara, T., and Yamashita, I. (1996). Genbank accession number D28954. Available at: www.ncbi.nlm.nih.gov/Genbank. Accessed on August 17, 2001.
Pakdel, F., Petit, F., Anglade, I., Kah, O., Delaunay, F., Bailhache, T., and Valotaire, Y. (1994). Overexpression of rainbow trout estrogen receptor domains in Escherichia coli: Characterization and utilization in the production of antibodies for immunoblotting and immunocytochemistry. Mol. Cell. Endocrinol. 104, 81–93.[Web of Science][Medline]
RxList (1999). Available at http://www.rxlist.com/99top.htm. Accessed August 17, 2000.
Scholz, S., and Gutzeit, H. O. (2000). 17--ethinylestradiol affects reproduction, sexual differentiation and aromatase gene expression of the medaka (Oryzias latipes). Aquat. Toxicol. 50, 363–373.[Web of Science][Medline]
Schwabl, H. (1993). Yolk is a source of maternal testosterone for developing birds. Proc. Nat. Acad. Sci. U.S.A. 90, 11446–11450.
Shioda, T., and Wakabayashi, M. (2000). Effect of certain chemicals on the reproduction of medaka (Oryzias latipes). Chemosphere 40, 239–243.[Medline]
Srivastava, R. K., and Van der Kraak, G. J. (1994). Insulin as an amplifier of gonadotropin action on steroid production: Mechanisms and sites of action on goldfish prematurational full-grown ovarian follicles. Gen. Comp. Endocrinol. 95, 60–70.[Web of Science][Medline]
Tchoudakova, A., Pathak, S., and Callard, G. V. (1999). Molecular cloning of an estrogen receptor-ß subtype from the goldfish, Carassius auratus. Gen. Comp. Endocriol. 113, 388–400.
Ternes, T. A., Stumpf, M., Mueller, J., Haberer, K., Wilken, R. D., and Servos, M. (1999). Behavior and occurrence of estrogens in municipal sewage treatment plants. I. Investigations in Germany, Canada and Brazil. Sci. Total Environ. 225, 81–90.
Thompson, S. C. (2000). Physiological indicators of endocrine disruptor exposure in Japanese medaka (Oryzias latipes): Relationship to reproduction and development. Master’s thesis, The University of Mississippi, University, MS.
Todo, T., Adachi, S., and Yamauchi, K. (1996). Molecular cloning and characterization of Japanese eel estrogen receptor cDNA. Mol. Cell. Endocrinol. 119, 37–45.[Web of Science][Medline]
Van der Kraak, G. J., Munkittrick, K. R., McMaster, M. E., Portt, C. B., and Chang, J. P. (1992). Exposure to bleached kraft pulp mill effluent disrupts the pituitary-gonadal axis of white sucker at multiple sites. Toxicol. Appl. Pharmacol. 115, 224–233.[Web of Science][Medline]
Yamamoto, T.-O. (1965). Estriol-induced XY females of the medaka (Oryzias latipes) and their progenies. Gen. Comp. Endocrinol. 5, 527–533.[Web of Science][Medline]
Zillioux, E. J., Johnson, I. C., Kiparissis, Y., Metcalfe, C. D., Wheat, J. V., Ward, S. G., and Liu, H. (2001). The sheepshead minnow as an in vivo model for endocrine disruption in marine teleosts: A partial life-cycle test with 17-ethinylestradiol. Environ. Toxicol. Chem. 20, 1968–1978.
CiteULike 
Connotea 
Del.icio.us What's this?
This article has been cited by other articles:
|
|
 |
|
  Z. Zhang and J. Hu Development and Validation of Endogenous Reference Genes for Expression Profiling of Medaka (Oryzias latipes) Exposed to Endocrine Disrupting Chemicals by Quantitative Real-Time RT-PCR Toxicol. Sci., February 1, 2007; 95(2): 356 - 368. [Abstract] [Full Text] [PDF]
|
|
|
|
 |
|
 M. D. Anway, C. Leathers, and M. K. Skinner Endocrine Disruptor Vinclozolin Induced Epigenetic Transgenerational Adult-Onset Disease Endocrinology, December 1, 2006; 147(12): 5515 - 5523. [Abstract] [Full Text] [PDF]
|
|
|
|
|
|
M. D. Anway and M. K. Skinner Epigenetic Transgenerational Actions of Endocrine Disruptors Endocrinology, June 1, 2006; 147(6): s43 - s49. [Abstract] [Full Text] [PDF]
|
|
|
|
 |
|
 M. D. Anway, A. S. Cupp, M. Uzumcu, and M. K. Skinner Epigenetic Transgenerational Actions of Endocrine Disruptors and Male Fertility Science, June 3, 2005; 308(5727): 1466 - 1469. [Abstract] [Full Text] [PDF]
|
|
|
|
 |
|
 T. Kristensen, E. Baatrup, and M. Bayley 17{alpha}-Ethinylestradiol Reduces the Competitive Reproductive Fitness of the Male Guppy (Poecilia reticulata) Biol Reprod, January 1, 2005; 72(1): 150 - 156. [Abstract] [Full Text] [PDF]
|
|
| ||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||||