* 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
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ABSTRACT |
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Key Words: teleost; vitellogenin; steroid; development; differentiation.
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INTRODUCTION |
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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.
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MATERIALS AND METHODS |
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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.
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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 (600010,000 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 (10,00013,000 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.
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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 510 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 Fishers 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.
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RESULTS |
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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; Fishers 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; Fishers 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).
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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; Fishers PLSD tests, p < 0.01) and increased significantly with adult reexposure at 8000 ng/l (Fishers 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
).
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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; Fishers 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; Fishers 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.
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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; Fishers PLSD tests, p < 0.05; Fig. 11).
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DISCUSSION |
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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 Thompsons 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.
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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.
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ACKNOWLEDGMENTS |
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NOTES |
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This research was supported in part by the U.S. Environmental Protection Agency. It has not been subjected to the Agencys peer and policy review and therefore does not necessarily reflect the views of the Agency. No official endorsement should be inferred.
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