ToxSci Advance Access originally published online on December 30, 2005
Toxicological Sciences 2006 90(2):490-499; doi:10.1093/toxsci/kfj085
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Inhibition of Follicular Development, Vitellogenesis, and Serum 17ß-Estradiol Concentrations in Zebrafish Following Chronic, Sublethal Dietary Exposure to 2,3,7,8-Tetrachlorodibenzo-p-Dioxin
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* Marine & Freshwater Biomedical Sciences Center, University of Wisconsin-Milwaukee, Milwaukee, Wisconsin 53204; Great Lakes WATER Institute, University of Wisconsin-Milwaukee, Milwaukee, Wisconsin 53204; and Department of Biological Sciences, University of Wisconsin-Milwaukee, Milwaukee, Wisconsin 53211
1 To whom correspondence should be addressed at Department of Biological Sciences, University of Wisconsin-Milwaukee, Milwaukee, WI 53211; E-mail: rjhutz{at}uwm.edu.
Received October 10, 2005; accepted December 20, 2005
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ABSTRACT |
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The environmental toxicant 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) is a potent endocrine disruptor with the ability to affect several biologic processes, including reproduction. In fish, sublethal exposure to TCDD is known to modulate overall reproductive capacity, but impacts on follicular development and vitellogenesis are unknown. Here we show that chronic, dietary exposure to 0.08, 0.32, or 0.80 ng TCDD female–1 day–1 decreased egg production by more than 50% and that spawning success was reduced by as much as 96%. Serum estradiol concentrations were decreased more than twofold, accounting, in part, for observed decreases in serum vitellogenin concentrations by as much as 29%. Our data suggest that decreased egg production is likely the result of TCDD-mediated inhibition of the transition from pre-vitellogenic stage follicles to vitellogenic stage follicles, as well as the induction of follicular atresia. The majority of reproductive toxicity of TCDD is likely due to direct impacts on the ovary, yet histopathologic observations suggest liver toxicity could also contribute to observed impacts on vitellogenesis. Importantly, even when overall egg production is not significantly affected, our data show that subtle physiologic changes induced by TCDD can lead to altered gonadogenesis. This suggests that long-term exposure to very low concentrations of TCDD could greatly affect fecundity and reproductive success in fishes.
Key Words: TCDD; follicular development; estrogen and vitellogenin concentrations; zebrafish.
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMETHODSRESULTSDISCUSSIONSUPPLEMENTARY DATAREFERENCES |
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Efforts to increase agricultural productivity through widespread use of insecticides, herbicides, fungicides, and pesticides, in addition to by-products of modern manufacturing processes, have resulted in the release of substantial amounts of toxicants into the environment. These chemicals have been proven to exert a profound effect on non-target species, and sublethal doses of toxic chemicals also influence the success of wildlife populations. Many of these environmental contaminants are capable of disrupting an animal's endocrine system, affecting development and reproduction (Ankley and Giesy, 1998
; Arcand-Hoy and Benson, 1998; Cooper and Kavlock, 1997; Tyler et al., 1998).
Polychlorinated dibenzodioxins are environmental contaminants known to adversely affect reproduction and early development in fish (Giesy et al., 1999; Munkittrick et al., 1991a, 1991b; Peterson et al., 1993; Poland and Knutson, 1982; Safe, 1990; Van der Kraak et al., 1992; Walker et al., 1996). The most toxic of these compounds is generally considered to be 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), which is a globally distributed, highly persistent compound with the potential to modulate several biologic processes that have an impact on growth and development. Although much progress has been made to delineate the toxic responses during early development, important questions remain, including potential sublethal effects on reproduction in adult fish. Studies indicate that exposure to high concentrations of TCDD can have a profound effect on the overall health and success of wild fish populations (Cooper and Kavlock, 1997; Malins and Ostrander, 1991). TCDD and TCDD-like compounds have been shown to decrease circulating 17ß-estradiol concentrations in rainbow trout (Hutz et al., 1999), white sucker (Munkittrick et al., 1991b; Van der Kraak et al., 1992), and Atlantic croaker (Thomas, 1990), which may be the result of direct impairment of the ovary or the hypothalamic–pituitary–gonadal (HPG) axis. Furthermore, a decrease in circulating 17ß-estradiol or alterations in follicle-stimulating hormone (FSH) and luteinizing hormone (LH) secretion could potentially affect vitellogenesis or oogenesis, resulting in a decrease in egg production and reproductive success.
Although the mechanisms by which TCDD exerts its toxicity have been extensively studied, much has yet to be learned about the direct effects of dioxin on the female reproductive system of fishes. The zebrafish provides a powerful vertebrate model system to investigate mechanisms by which TCDD affects the reproductive system of some fish, and findings will be applicable to other fishes, as well as to humans. We have shown previously that chronic, sublethal exposure to TCDD affects reproductive success of zebrafish by interfering with reproductive effort (ovary somatic index, OSI) and survival of offspring (King Heiden et al., 2005a). Follicular development and vitellogenesis, secretion of gonadotropins and steroid hormones, and steroidogenesis all play important roles in reproductive success, and are all potentially affected by TCDD. This study was designed to investigate the impact of sublethal dietary TCDD exposure on egg production, 17ß-estradiol concentrations, and alteration of follicular development and vitellogenesis in female zebrafish in an effort to further develop the zebrafish as a model to study the signaling pathways responsible for dioxin toxicity.
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METHODS |
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Experimental animals.
Adult female (AB strain, Zebrafish International Resource Center) and male (golden leopard strain, Ekwill Farms) zebrafish were housed separately and acclimated for several weeks prior to the initiation of experiments. Fish used in this study were 6 months old, and were known to be consistent spawners. Different strains were used to facilitate differentiation between sexes, as zebrafish do not show prominent signs of sexual dimorphism. Fish were maintained at 26°–28°C on a 14-h light and 10-h dark cycle in a flow-through buffered, dechlorinated water system at the UWM Marine and Freshwater Biomedical Sciences Center (Milwaukee, WI). All experimental procedures were approved by the University of Wisconsin-Milwaukee Animal Care and Use Committee.
Food preparation.
TCDD was synthesized and purified to >99% by the manufacturer (Cambridge Isotope Laboratories, Inc., Andover, MA). Food yielding a final concentration of 0, 10, 40, or 100 ng TCDD/g food (ppb) was prepared as previously described (King Heiden et al., 2005a). Briefly, the TCDD stock solution was diluted in acetone, added to trout chow (Zeigler, Gardner, PA), and swirled to ensure homogeneous distribution, after which the acetone was evaporated from the food in a fume hood overnight.
Exposure regimen and experimental design.
Experiments were conducted in two phases: pre-exposure (baseline) and exposure, and were initiated with 26 females in each exposure group. During the pre-exposure phase, all fish were fed brine shrimp nauplii and trout chow daily for 3 weeks. During the exposure phase, females were fed (en masse) the trout chow diet containing 0 (acetone only), 10, 40, or 100 ng of TCDD/g food (ppb) and brine shrimp nauplii 5 days a week for a period of 4 weeks. To be consistent with our previous work (King Heiden et al., 2005a), 2 days each week, fish were fed trout chow that did not contain TCDD. During the exposure phase, males received the same diet as during the pre-exposure phase. Fish were fed to satiety and, based on the average food consumed per fish, females received an estimated applied dose of 0 (vehicle control), 0.08, 0.32, or 0.80 ng TCDD/female/day. Food consumption by individual fish was not controlled; however, over the course of the experiment, fish did not show significant change in weight within or between treatment groups (within 2 standard deviations of the mean), suggesting that all fish received similar amounts of food, and therefore a similar dose within treatment groups. We have previously demonstrated that, with this experimental regimen, total body burdens increased over time in a dose-dependent manner (King Heiden et al., 2005a). Estimated accumulation rates were 0.04, 0.33, and 0.77 ng TCDD female–1 week–1 for the 10, 40, and 100 ppb treatment groups, respectively. A positive linear relationship exists between estimated applied dose and measured whole-body concentrations (y = 0.31x + 0.04, r2 = 0.93) and mean net dietary assimilation was shown to be approximately 96%. Previously measured total body burdens over time are listed in Table 1 for reference.
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Impacts on spawning and egg production.
During both phases of the experiment, mortality and general health were monitored daily and females were spawned with untreated males once weekly following 5, 10, 15, and 20 days of dietary exposure. For spawning, females were randomly allocated to one of six spawn tanks in the evening, and males were randomly added to tanks 20–30 min prior to simulated dawn in a ratio of 2 females:1 male. Upon completion of spawning (approximately 1.5 h), one female from each spawn tank was removed for further analyses as described below; remaining females and males were returned to their respective holding tanks. Eggs were then collected from spawned females and maintained at 28°C until counted. Eggs were observed under a dissecting microscope approximately 3–4 h post-fertilization to determine whether they had been fertilized; both fertilized and unfertilized eggs were counted. Spawning was considered successful if the group of females together produced
10 eggs per female. If average egg production was <10 eggs/female, then spawning of all females within that spawn tank was considered unsuccessful; spawning success was calculated based on the percent of females within each treatment group that successfully spawned.
Impacts on serum 17ß-estradiol and vitellogenin concentrations.
Following 5, 10, 15, and 20 days of dietary exposure, six females from each treatment group were anesthetized by submersion in 0.1 g/l 3-aminobenzoic acid ethyl ester (MS-222, Sigma) following spawning, and wet weight and total length were recorded for each fish. Blood was collected from the caudal vein using a heparinized capillary tube, centrifuged (3000 x g), and plasma was removed and stored at –80°C. Serum 17ß-estradiol (E2) and vitellogenin (Vtg) concentrations were determined using ELISA according to the manufacturer's instructions (Cayman Chemical). For E2 analysis, plasma samples were diluted 1:50 with sample buffer, and then diluted further 1:144,000 with vitellogenin sample buffer for Vtg analysis. A few samples did not contain detectable E2 concentrations and were therefore assigned the detection limit as its concentration (1.4 pg/ml). All samples had detectable Vtg concentrations.
Histopathology.
Following 20 days dietary exposure, females were processed for histolomorphometric analysis. The heads and tails were removed, and an incision was made along the ventral surface. The samples were fixed in Altmanns solution (Humason, 1972) and decalcified as described by Bauer and Goetz (2001). Specimens were dehydrated in a graded series of ethanol, cleared in xylene, and embedded in paraffin. Samples were sectioned in coronal orientation at 7–10 µm, mounted on slides, and stained with hematoxylin and eosin. Histopathology was performed blind to the observer and was characterized according to van der Ven et al. (2003). Sectioning was standardized according to the relative position of the organ of interest to other structures, such that sections analyzed showed both lobes of the ovary, liver, and swim bladder simultaneously. Two separate sections from the same general area were assessed for each female (n = 4–6 for each treatment group).
To assess impacts on ovarian development, digital micrographs at 12x magnification were analyzed using Image-Pro Express software (Media Cybernetics). Oocytes were classified according to Nagahama (1983), Selman et al. (1993), and Miranda et al. (1999). Follicles were classified as atretic if granulosa cells were reduced in number, were in close apposition to atretic follicles, and were hypertrophied and filled with yolk (as described by Miranda et al., 1999). Follicles in the early perinucleolus stage (<0.002 mm2) were counted but not measured. All other follicles were counted, measured, and classified as being in one of the following three general categories:
- "Primary growth" (0.002–0.015 mm2; includes follicles in the late perinucleolus stage with the vitelline envelope just forming and showing no yolk bodies);
- "Secondary growth" (0.015–0.09 mm2; includes follicles in the oil droplet stage, with irregular shape, cuboidal follicle cells, vitelline envelope, and apparent oil droplets or yolk vesicles); or
- "Vitellogenic" (0.09–0.40 mm2; follicles in the primary, secondary or tertiary yolk stage with crystalline yolk globules present, germinal vesicles were not always apparent; also includes mature follicles that were not atretic).
), the total number of follicles and the numbers within different follicle classes were normalized to the total area observed. Liver histopathology was determined based on changes compared vehicle control. Liver tissue was observed at 50x magnification, and scored for each toxic endpoint (lipidosis, hepatocyte hypertrophy, and number of hepatocyte nuclei) according to severity. Sections were given a severity score of zero if no histopathology was observed, one if histopathology was mild (present less than 25% of section), two if histopathology was moderate (present 25–75% of section), or three if severe histopathology was observed (present in more than 75% of section). Scores for these signs of liver toxicity were then averaged to give an overall liver toxicity score.
Data analysis.
Statistical analysis of the data was performed using Sigma-Stat 2.0 (SPSS, Inc) and presented as means ± standard error of the mean (SEM). Since egg production can be variable across individual females, one-way analysis of variance (ANOVA) was used to detect treatment-related effects on egg production compared with egg production of the females within respective treatment groups during the pre-exposure phase of the experiment. Two-way ANOVA was used to detect treatment-related effects on serum 17ß-estradiol and vitellogenin concentrations with respect to "TCDD dose versus time." One-way ANOVA was used to detect treatment-related effects on each histopathologic parameter. Data were evaluated for homoscedasticity (Levene median test) and for normality prior to ANOVA. Pair-wise multiple comparisons were conducted using the Tukey test, with significant differences identified at p < 0.05. Linear regression was performed to establish relationships between estradiol and vitellogenin concentrations, and between average estradiol and vitellogenin concentrations following 20 days of dietary exposure to TCDD and the average proportion of vitellogenic follicles. If an exact p value is not given, p < 0.05 is considered to be statistically significant.
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RESULTS |
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Egg Production
Dietary exposure to TCDD did not impact fertilization of eggs (data not shown). While females in the 10-ppb treatment group showed a significant decrease in egg production after five days of dietary exposure (p = 0.005); overall they did not show reduced egg production compared with baseline, and spawning success was not affected (Fig. 1A and 1B). Females in the 40- and 100-ppb treatment groups showed a significant decrease in egg production, producing at least 50% fewer eggs compared to baseline (p = 0.04 and < 0.001, respectively) (Fig. 1A). Spawning success was also decreased, with fewer than 70% of the females spawning successfully (Fig. 1B). Overall egg production by females in the 40-ppb treatment group was decreased; however, the females that spawned successfully did not show a significant reduction in egg production (Fig. 2A). In contrast, the females in the 100-ppb treatment group that spawned successfully showed a significant reduction in egg production of greater than 50% (p < 0.001) (Fig. 2B).
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Serum 17ß-Estradiol Concentrations
Serum 17ß-estradiol (E2) concentrations of females treated with food containing 40 and 100 ppb TCDD for 5 days were significantly decreased by more than 36% compared with control, whereas E2 concentrations were significantly decreased by more than 50% in all treatment groups following 10 and 15 days of dietary exposure (Fig. 3A). Following 20 days of dietary exposure, females in the 10-ppb treatment group showed a 52% reduction in serum E2 concentrations compared with control, and females in the 40- and 100-ppb treatment groups showed significantly lower serum E2 levels (72% and 84% reductions, respectively) (Fig. 3A).
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Serum Vitellogenin Concentrations
Although TCDD affected vitellogenin (Vtg) concentrations over time, effects were not dose dependent (Fig. 3B). After 5 days of dietary exposure, Vtg concentrations were significantly reduced by 12% compared with control, and after 10 days of dietary exposure Vtg concentrations were reduced by 15–20%. Following 15 and 20 days of dietary exposure, vitellogenin concentrations were significantly reduced by approximately 28% (Fig. 3B). As expected, there was a positive relationship between reduced E2 concentrations and reduced Vtg concentrations (r2 = 0.56; p < 0.001, data not shown).
Ovarian Development
Chronic dietary exposure to sublethal concentrations of TCDD affected ovarian development (representative micrographs are shown in Figure 4 A–D). The total number of follicles was significantly reduced following 20 days of dietary exposure to food containing 100 ppb TCDD, and the number of atretic follicles was significantly increased in a dose-dependent manner (Table 2). Approximately 40% of the follicles from females receiving vehicle control were primary growth follicles, 40% were secondary growth follicles, and the remaining 20% were vitellogenic follicles (Fig. 5). There was no impact on the percentage of follicles in the primary growth phase of development in treated females compared with control. Ovaries from treated females were comprised of primarily secondary growth follicles (approximately 50%), and had significantly fewer vitellogenic follicles compared with control (11–6%; Fig. 5). TCDD exposure affected the growth of secondary growth follicles, with females from the 40- and 100-ppb treatment groups having significantly smaller (8–10%) secondary growth follicles compared with control (Table 3). Only females from the 10-ppb treatment group had smaller (12%) vitellogenic follicles (Table 3). As expected, there was significant regression between the mean proportion of vitellogenic and mean serum E2 concentrations (r2 = 0.97; p = 0.017, data not shown); however, the average proportion of vitellogenic follicles could not be predicted by vitellogenin concentrations (r2 = 0.85; p = 0.077, data not shown).
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Liver Toxicity
TCDD-induced liver toxicity manifested itself in a dose-dependent manner (Table 4) as described by Zodrow et al. (2004)
and Spitsbergen et al. (1988a; 1988b). Livers from females exposed to TCDD exhibited lipidosis, hepatocyte hypertrophy, and an overall decrease in the number of hepatic nuclei (Fig. 4E and 4F).
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DISCUSSION |
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TCDD is a reproductive toxicant and endocrine disruptor in nearly all vertebrates, and fish are among the most sensitive vertebrates to its toxic effects. Several studies have shown that sublethal exposure to TCDD affects overall reproductive capacity of fish (Giesy et al., 2002
; King Heiden et al., 2005a; Tietge et al., 1998; Walker and Peterson, 1994); nevertheless, the mechanisms by which TCDD induces these reproductive alterations have not been well characterized. Particularly lacking is information regarding the effects of sublethal TCDD exposure on piscine follicular development. Here we show that TCDD-mediated impacts on egg production were likely a result of impaired follicular development, as well as induced ovarian toxicity resulting in the increase in the number of atretic follicles. Furthermore, even when overall egg production was not significantly impacted, subtle physiologic changes induced by TCDD led to altered follicular development, suggesting that long-term exposure to very low concentrations of TCDD could greatly affect fecundity.
Based on our previous work (King Heiden et al., 2005a) approximate accumulation of 0.5–1 ng TCDD/g female had little to no effect on overall egg production or spawning activity. However, follicular development and atresia, serum 17ß-estradiol, and vitellogenin concentrations, as well as growth of vitellogenic follicles were affected in these females, suggesting that continued exposure to such levels of TCDD could exert an effect on egg production. An estimated accumulation of 3–7 ng TCDD/g female reduced egg production by greater than 50%, and only 70–80% of females were able to spawn successfully, whereas accumulation of 13–15 ng TCDD/g reduced egg production by 96%, with only 20–40% of females capable of spawning. This impact on egg production was also reflected in the impact on follicular development and atresia. It is interesting that the females in the 40 ppb treatment group that spawned successfully did not show reduced egg production, while those in the 100 ppb treatment group that spawned successfully produced approximately 50% fewer eggs after an estimated accumulation of similar concentrations of TCDD (Fig. 2; Table 1). Perhaps then, the rate by which TCDD accumulated within females exerted an effect on ovarian toxicity and egg production. Nevertheless, overall egg production was decreased, and our data suggest that fecundity can be affected by chronic exposure to sublethal concentrations of TCDD.
Several studies have shown that acute exposure to TCDD reduces circulating 17ß estradiol (E2) concentrations in mammals (Chaffin et al., 1996; Roby, 2001; Safe, 1995); thus it is not surprising that the same holds true for fish. Both carp and white sucker exposed to effluents containing elevated levels of TCDD showed reduced serum E2 concentrations (Munkittrick et al., 1991b, 1998; Sakamoto et al., 2003; Van der Kraak et al., 1992). In rainbow trout, the AHR-agonist beta-naphthoflavone diminished serum E2 concentrations in both a time-dependent and dose-dependent manner (Hutz et. al., 1999), and E2 concentrations of Atlantic croaker decreased when exposed to similar compounds (Thomas, 1990). In zebrafish, an estimated accumulation of as little as 1 ng TCDD/g female resulted in a 1.6-fold decrease in serum E2 concentrations compared with control, and E2 concentrations were decreased 2-, 4-, and 6-fold following an estimated accumulation of 1, 7, and 15 ng TCDD/g female, respectively. In fish, E2 produced by the developing follicle stimulates the liver to produce the yolk protein vitellogenin, which is sequestered and processed by the developing oocyte to provide nutrients for offspring. The observed decrease in serum E2 concentrations following TCDD exposure likely accounts for at least part of the decreased serum vitellogenin concentrations. However, the females in the 10-ppb treatment group did not show significantly lower E2 concentrations after 5 days of dietary exposure (but serum vitellogenin concentrations were reduced by approximately 12% compared with control), so toxicity in the liver may have also been a contributing factor. This is supported by in vitro studies indicating that TCDD impairs vitellogenesis in fish hepatocytes (Anderson et al., 1996; Bemanian et al., 2004; Smeets et al., 1999a; 1999b). Also, the addition of vitellogenin to developing oocytes accounts for a significant portion of the overall weight of the ovary. The observed reduction in serum E2 and vitellogenin concentrations likely contributed to the previously described decrease in ovary somatic index (King Heiden et al., 2005a), which has also been described by others (Sakamoto et al., 2003).
17ß-Estradiol plays a critical role in piscine ovarian development, particularly by inducing the synthesis of vitellogenin and by mediating pituitary gonadotropin control of oogenesis (Nagahama et al., 1995; Peter and Yu, 1997; Wallace and Selman, 1978; Young et al., 2005). Follicular development in zebrafish was attenuated by TCDD exposure, as has been described in mammals (Gao et al., 1999a, 1999b; Heimler et al., 1998a, 1998b; Petroff et al., 2000). Our work suggests that chronic, sublethal exposure to TCDD disrupted the transition from secondary growth stage to vitellogenic follicles and induced follicular atresia, as was reported to occur in zebrafish after acute exposure to concentrations of TCDD that induced overt toxicity (Wannemacher et al., 1992). And while impacts on follicular atresia were not described, Sakamoto et al. (2003) reported a reduction in the number of vitellogenic follicles in carp living in TCDD-contaminated effluents. Inhibitory effects of TCDD on estrogen biosynthesis likely contributed to the observed effects on follicle development and vitellogenesis in zebrafish.
While it is clear that TCDD exposure results in decreased E2 concentrations and impaired follicular development, the mechanism(s) that underlie these toxicities is complicated and is far from being understood. The best-characterized mechanism for TCDD action is via the cytoplasmic aryl hydrocarbon receptor (AHR) (Safe, 1986). Once bound by ligand (e.g., TCDD), the complex is translocated to the nucleus, where it dimerizes with AHR nuclear transporter protein (ARNT). The ligand-bound AHR/ARNT heterodimer then interacts with AHR responsive elements (AHREs), and can either activate or suppress the transcription of several genes, primarily metabolizing enzymes or growth-regulatory proteins. One of the most extensively studied AHR-inducible transcripts is cytochrome P4501A1 (CYP1A1). In addition to metabolizing a large number of toxic compounds, it also plays an important role in E2 metabolism (Zhu and Conney, 1998). TCDD-induced transcription of CYP1A1 could then contribute to the observed decrease in 17ß-estradiol concentrations by enhancing estrogen metabolism, as has been suggested by others (Pocar et al., 2003, 2005; Safe and Krishnan, 1995). Also, studies with mammals have shown that TCDD impairs aromatase mRNA expression and activity (Dasmahapatra et al., 2000). This enzyme is responsible for converting androgens (testosterone) to estrogens (17ß-estradiol), and its activity increases during the period of vitellogenesis in fishes (Nagahama et al., 1995; Young et al., 1983). TCDD could likewise inhibit aromatase expression or activity in fish, resulting in reduced 17ß-estradiol synthesis. Preliminary experiments performed in our laboratories using the same experimental design described here demonstrate that Cyp19 mRNA expression (specifically Cyp19a1) in the ovary is in fact reduced 1.5-fold and 2.0-fold after 15 days of dietary exposure to food containing 40 and 100 ppb TCDD, respectively (King Heiden et al., 2005b).
In addition to stimulating vitellogenesis, E2 can exert a positive feedback on the hypothalamic-pituitary-gonadal axis stimulating the release of FSH (important for gametogenesis and steroidogenesis), and high concentrations of E2 are necessary for the induction of the preovulatory LH surge, resulting in the maturation of oocytes (Kobayashi et al., 1989; Nagahama et al., 1995; Peter and Yu, 1997; Tyler et al., 1991). TCDD exposure resulting in reduced E2 concentrations could then diminish this positive feedback, resulting in reduced gonadotropin secretion and/or signaling and thereby affect ovarian development, as has been suggested to occur in rats (Gao et al., 2001; Petroff et al., 2001). Alternatively, TCDD could directly inhibit ovarian development by altering the expression of genes important for achieving oocyte competence, consequently reducing E2 and Vtg concentrations. In fish, peptide growth factors including activins, transforming growth factors (TGF
and TGFß), and epidermal growth factors (Ge, 2005) are important in the regulation of follicle development, oocyte maturation, and vitellogenesis, and are therefore promising candidate genes that could be dysregulated by TCDD. Direct evidence that disrupted expression of such genes is responsible for TCDD-induced ovarian toxicity is lacking, but recent work in rat and human suggests that it plays a role (Dohr et al., 1995; Gao et al., 1999a; Hirakawa et al., 2000a, 2000b; Li et al., 1995; Mizuyachi et al., 2002; Sutter et al., 1991; Yang et al., 1999; Yang, 1999).
In conclusion, this study demonstrates that chronic, sublethal exposure to TCDD disrupts reproductive physiology, specifically serum 17ß-estradiol and vitellogenin concentrations, follicular development, and vitellogenesis in female zebrafish. TCDD likely affects follicular development through a combination of depressed capacity for ovarian steroid synthesis, diminished gonadotropin production/action, and altered steroid metabolism, suggesting that TCDD can affect the ovary directly, as it modulates the pituitary-gonadal axis (Gagnon et al., 1994; Hutz, 1999; Van der Kraak et al., 1992). For example, TCDD could disrupt critical cellular signals that regulate the transition from the previtellogenic to vitellogenic follicle via AHRE-mediated alterations in gene transcription, thereby contributing to the observed decrease in ovarian development and reduced reproductive capacity. To better understand the mechanisms by which TCDD exerts its reproductive toxicity, additional studies are clearly needed to better characterize the specific cellular pathways that are activated or inhibited by TCDD. The zebrafish model system provides an ideal system within which to further examine these mechanisms.
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Supplementary data are available online at http://toxsci.oxfordjournals.org/.
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ACKNOWLEDGMENTS |
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We gratefully acknowledge Angela Schmoldt, Peggy Biga, and especially Marlies Rise, for their assistance with sectioning material for histopathologic analyses. We also thank Frederick Goetz for helpful and stimulating discussions of our data and for reading of this manuscript. The Zebrafish International Resource Center (ZIRC) provided the fish used in these experiments. ZIRC is supported by NIH-NCR (P40 RR12546). This work was supported in part by the Environmental Protection Agency (T.K.H., GRO MA916290), the National Institutes of Health (R.J.H., ES011569 [GenBank] ), a Shaw Scientist Award (M.J.C.) from the Greater Milwaukee Foundation, the Midwest Regional Chapter of the Society of Toxicology (Young Investigator Award to T.K.H.), and the UW-Milwaukee NIEHS Marine and Freshwater Biomedical Sciences Center (ES04181).
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