Nonylphenol Affects Gonadotropin Levels in the Pituitary Gland and

Jun 14, 2001 - Pituitary gland luteinizing hormone (LH) content was significantly lower in fish exposed to 85.6 μgNP/L, and LH gene expression was ...
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Environ. Sci. Technol. 2001, 35, 2909-2916

Nonylphenol Affects Gonadotropin Levels in the Pituitary Gland and Plasma of Female Rainbow Trout C A T H E R I N E A . H A R R I S , * ,† EDUARDA M. SANTOS,‡ AFSANEH JANBAKHSH,§ TOM G. POTTINGER,| CHARLES R. TYLER,‡ AND JOHN P. SUMPTER† Department of Biological Sciences, Brunel University, Uxbridge, Middlesex, UB8 3PH, U.K., School of Biological Sciences, The Hatherly Laboratories, University of Exeter, Prince of Wales Road, Exeter, EX4 4PS, U.K., Scientific Services Department, Essex and Suffolk Water Company, Hall Street, Chelmsford, Essex, CM2 0HH, U.K., and Centre for Ecology and Hydrology, Natural Environment Research Council, Windermere Laboratory, Ambleside, Cumbria, LA22 0LP, U.K.

Female rainbow trout (Oncorhynchus mykiss) were exposed to 4-nonylphenol (NP) at (mean measured) concentrations of 0.7, 8.3, and 85.6 µg/L, for 18 weeks, during early ovarian development. Fish were sampled sublethally every six weeks, and terminal samples were taken at 18 weeks. NP induced an estrogenic effect (the synthesis of vitellogenin) at concentrations of 8.3 and 85.6 µg/ L. An effect on gonadotropin synthesis and secretion was also observed. Plasma follicle stimulating hormone (FSH) levels and FSH gene expression in the pituitary were the most sensitive endpoints assessed, being reduced at the lowest dose employed (0.7 µgNP/L). Pituitary gland luteinizing hormone (LH) content was significantly lower in fish exposed to 85.6 µgNP/L, and LH gene expression was suppressed in fish exposed to 8.3 and 85.6 µgNP/L. In contrast, plasma LH concentration increased in these fish, but by a very minor absolute amount, and returned to control levels by the final sampling time. Gonadal development ceased in the fish exposed to 85.6 µgNP/L, and steroidogenesis in these fish was also markedly inhibited. Although the mechanisms underlying these responses are unknown, this study demonstrates that NP has adverse effects on pituitary function that can result in inhibition of ovarian development.

Introduction Reproduction in fish, as in all vertebrates, is ultimately controlled by the brain, via hormones from the pituitary gland (the gonadotropins (GTH)) which control gonadal development in both sexes. In fish, two gonadotropins have been identified (1, 2) which were originally referred to as GTH I and GTH II, but more recently have been found to be similar * Corresponding author phone: +44-1895-274000; fax: +44-1895274348; e-mail: [email protected]. † Brunel University. ‡ University of Exeter. § Essex and Suffolk Water Co. | Centre for Ecology and Hydrology. 10.1021/es0002619 CCC: $20.00 Published on Web 06/14/2001

 2001 American Chemical Society

to tetrapod FSH and LH, respectively (3). It is now considered that GTH I and II play roles in fish similar to those played by FSH and LH, respectively, in mammals; this is reflected in the seasonal profiles of plasma FSH and LH concentrations in the female rainbow trout (4) which mimic those in the oestrus cycle in mammals. GTH I and II are therefore referred to as FSH and LH, respectively, in this paper. This axis, the so-called brain-pituitary-gonadal (BPG) axis, is in turn regulated by feedback loops. 17β-Estradiol (E2) and aromatizable androgens have been shown to have a positive feedback action on LH synthesis and/or secretion in immature fish (5-13). Fewer data are available concerning the response of FSH to steroids in fish. Available data on FSH, however, unanimously support a negative feedback response of FSH synthesis and/or secretion to steroid hormones (12-15). Over recent years, 4-nonylphenol (NP) has become well established as an “environmental estrogen”. NP is a widely employed industrial chemical which is commonly found in river systems, often at concentrations which have been shown to induce estrogenic effects in fish (e.g., Jobling et al. (16)). 4-Nonylphenol has a variety of applications, however, the majority of NP is used in the production of nonylphenol polyethoxylates (NPEOs), which are widely used nonionic surfactants. NPEOs are constituents of both domestic and industrial detergents which enter the sewage system where they are degraded into the smaller chain ethoxylates and alkylphenols before their release to the aquatic environment. Essentially, NP contamination tends to be concentrated around urban or industrialized areas, and in the wider aquatic environment it is generally detected at concentrations of less than 10 µg/L. Nonetheless, concentrations of up to 180 µg/L have been reported in the River Aire (17). To date, the effects of environmental estrogens (EEs) in fish have been investigated at the level of the final target organ (e.g., vitellogenin synthesis in the liver; intersexuality in the gonad), and data reporting the effects of EEs at higher levels of the BPG axis are limited (but see refs 18-20 for some recent data indicating effects of NP on gonadotropins in fish, and Khan and Thomas (21) for a report of the effect of o,p′-DDT on LH in fish). Estrogen receptors (ERs) occur in the brain (hypothalamus) and pituitary gland, where they play important roles in controlling LH and FSH synthesis and secretion, and it therefore seems likely that estrogen mimics, such as NP, could also have effects at the higher levels of the BPG axis. This study was undertaken to determine whether NP does have effects at such higher levels.

Methods Experimental Design. Fish were exposed to nominal concentrations of 1, 10, and 100 µgNP/L. A solvent control tank was also set up, in which fish were exposed to methanol (MeOH) at the concentration present in the NP-treated tanks (0.002%). This concentration was well within the level suggested by the U.S. Environmental Protection Agency (U.S. EPA) for use of solvents in aquatic toxicity test systems (0.01%, (22)). A dilution water tank was run as an absolute control. The experiment was conducted at the NERC Centre for Ecology and Hydrology and was carried out in 1500-liter circular glass-fiber tanks, each supplied with a constant flow of lake water (20 L/min), at ambient temperature (increasing from 5 to 18 °C during the experiment). The tanks and delivery system were equilibrated with NP for three weeks prior to the introduction of the fish. NP (99% purity, and consisting of a mixture of isomers) was purchased from Acros Organics, VOL. 35, NO. 14, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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(Loughborough, Leics, U.K.), and was administered to the tanks at a rate of 0.4 mL/min. Female rainbow trout (Oncorhynchus mykiss), 2+ years in age and with a mean weight of 0.598 kg, were obtained from New Mills Trout Farm (Brampton, Cumbria, U.K.). Only females were studied because they have higher levels of GTH than males (4), thus facilitating investigations on the effects of a xenoestrogen on GTHs. Thirty randomly selected fish were placed in each tank, and pre-dose control samples were collected from an additional thirty fish. During the course of the study, the fish were fed three times per week with a commercial feed (Trouw Aquaculture Standard Expanded) at the manufacturer’s recommended rate. The experiment was initiated in early March, at a time prior to the expected increase in plasma FSH concentration that accompanies the onset of vitellogenin uptake into oocytes. Fish were sampled every six weeks, and the experiment was terminated in midJuly, when FSH levels were expected to be at their peak (see Prat et al. (4) for details of changes in FSH and LH concentrations during the reproductive cycle in rainbow trout). Extraction and Analysis of Water Samples. Water samples were collected in solvent-washed glass bottles, at six-week intervals (at the same time as the biological samples); one sample was collected per treatment at each timepoint. All samples were preserved with 10 mL/L of 38% formaldehyde and stored in the dark at 4 °C prior to extraction. Samples were concentrated using sep-pak cartridges containing a packing material of 200 mg of ethyl bonded silica (C2; JT Bakers, Berkshire, U.K.), according to Janbakhsh (23). The final extract was redissolved in 250 µL of methyl-tert-butylether (MTBE; BDH, Poole, U.K.), which contained 4 mg/L 2-phenylphenol (Aldrich, Gillingham, U.K.) as an internal standard. Extracts were analyzed by normal-phase HPLC and fluorescence detection. The conditions used have been described in Janbakhsh (23). The limit of detection of NP was calculated according to the Manual on Analytical Quality Control for the Water Industry (24) and was found to be 0.242 µg/L. Recoveries were calculated individually for each treatment at each sampling point; final sample concentrations were corrected for recovery. Biological Analyses. Fish were anesthetized using 2-phenoxyethanol (1:2000). Blood was sampled from the caudal sinus using heparinized syringes, treated with aprotinin (2 TIU/mL), and kept on ice prior to centrifugation. Blood plasma was then drawn off and frozen at -20 °C until use. Weight and fork length were also recorded; the coefficient of condition of the fish was calculated using the formula [(weight/length3) × 100]. At the terminal sample (18 weeks), ovaries and livers were removed and weighed, and pituitaries were collected and frozen in liquid nitrogen until transfer to -80 °C in the laboratory. The gonadosomatic index (GSI) and hepatosomatic index (HSI) were calculated, providing indications of gonad and liver weights, respectively, as a percentage of total body weight. FSH, LH, vitellogenin (VTG), E2, and testosterone concentrations were assessed by radioimmunoassay (RIA), according to Prat et al. (4) (LH); Santos et al. (25) (FSH); Sumpter et al. (26) (VTG); and Carragher et al. (27) (E2 and testosterone). Pituitary glands were extracted using a method described by Hassin et al. (28), whereby a single pituitary could be used to provide samples for both GTH protein assay by RIA, and GTH subunit mRNA analyses. Frozen pituitary glands were homogenized in 200 µL of LiCl (3 M)/urea (6 M). Aliquots (10 µL) were removed and diluted with 990 µL of protein assay buffer prior to analysis by radioimmunoassay. To the remaining homogenate, a 0.1 volume of 2 M sodium acetate and 2.5 volumes of ethanol were added. The sample was 2910

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stored at -70 °C for at least 2 h prior to centrifugation at 14000 rpm for 10 min. The resulting pellet was resuspended in 700 µL of TES (10 mM TRIS, 1mM EDTA, 0.5% SDS) and subsequently extracted twice with 700 µL of acid-phenol/ chloroform (50:50). The final aqueous phase was transferred to a fresh Eppendorf tube, the total RNA in the sample was precipitated as described above, and the sample was stored at -70 °C for at least 2 h. Following centrifugation, this precipitate was resuspended in 40 µL of DEPC-treated (i.e., RNase-free) water. Dot Blot Hybridization. Chum salmon LH cDNA (29) was a gift from Dr. F. Le Gac (Laboratoire de Physiologie des Poissons, Institute National de la Recherche Agronomique, France), and its use had been previously validated for detection of rainbow trout LH mRNA by Weil et al. (30) and Gomez et al. (31). Trout β-actin cDNA (32) was also a gift from Dr. F. Le Gac. A fragment (around 500 base pairs in size) of rainbow trout FSH cDNA containing the coding region was isolated and cloned from rainbow trout pituitary glands. A PCR reaction was undertaken with reverse-transcribed pituitary cDNA, using primers designed to lie on either side of the targeted coding region. The sequences of the primers were as follows: forward primer, 5′-GCGATAGCACATCAATGGAAA-3′; reverse primer, 5′-GATTCCTGAATAGACCTGTTCT-3′. The resulting PCR product was cloned into the pGEMT-Easy vector. Of the target sequence, 185 nucleotides were sequenced, and this sequence was found to be 88% homologous to the corresponding region of chum salmon FSH (Figure 1). Probes were labeled using a random primer system (Amersham Pharmacia Biotech, Bucks, U.K.) with 5′-[a-32P]dCTP (3000 Ci/mM), just prior to hybridization. Amounts of total RNA in the extracted samples were quantified by Gene Quant (Pharmacia), and 10- and 5-µg total RNA samples were loaded onto nylon membranes (Hybond N+, Amersham Pharmacia Biotech, Bucks, U.K.). Membranes were hybridized with three successive cDNA probes: FSH, LH, and β-actin. Membranes were prehybridized at 42 °C for 3 h in hybridization buffer (50% formamide, 5× SSC, 5× Denhardts reagent (33), 1% SDS, 0.1 g/mL dextran sulfate, and 0.2 M phosphate buffer), and 100 µg of calfthymus DNA/mL buffer (used to suppress nonspecific hybridization of the probes); hybridization was carried out overnight, using the same buffer. Membranes were washed, and the radioactive signal was quantified using a Storm phosphorimager (Molecular Dynamics. Inc.), which was linked to an ImageQuant software package (Molecular Dynamics. Inc.). Statistical Analysis. The data were log-transformed where necessary, and analyzed by either multi-factorial ANOVA or one-way ANOVA. Fisher’s PLSD was used as a post-hoc test for most data sets. The Bonferroni-Dunn test - a more stringent post-hoc test - was applied to the plasma FSH data, because of the large number of outlyers in this dataset. Regression analyses were also undertaken to establish the presence (or otherwise) of dose-related responses. The statistical package used was SUPERANOVA (Abacus Concepts, Berkeley, CA).

Results and Discussion To date, most reports on the response of fish to NP (and other xenoestrogens) have relied upon fairly brief exposure periods (a few weeks). In the experiment reported here, the exposure was maintained for 18 weeks. Experiments involving long-term exposure (and, ultimately, multi-generation studies) are required in order to more fully understand the effects that fish may exhibit following chronic, life-long environmental exposure to such chemicals. In addition, our study employed large numbers of fish in each treatment tank (n ) 30 in most cases). Such high numbers increase the likelihood of establishing subtle endocrinological changes.

FIGURE 1. Comparison between the sequenced region of the isolated rainbow trout FSH cDNA (top) and the corresponding region of chum salmon cDNA (bottom). The high homology between the two nucleotide sequences can be clearly seen.

FIGURE 2. Coefficient of condition of the fish throughout the experiment. The data derived from the fish exposed to NP were compared with those obtained from the fish exposed to solvent-spiked tank water, which is considered to be the most appropriate control. Concentrations of NP in the Tank Water. In most cases, the actual concentrations of NP in the tanks were maintained close to nominal concentrations; therefore, a 10-fold differential in NP concentration was maintained between treatments. The measured concentrations were, as is often the case, slightly lower than nominal concentrations, with the mean measured concentrations (( standard error) being 0.7 ( 0.4, 8.3 ( 0.9, and 85.6 ( 2.7 µgNP/L (n ) 4 in all treatments). Effects of NP on Somatic Growth and Reproductive Endpoints. In all but the highest treatment, the number of fish sampled at each timepoint was g28 (having omitted any males from the analyses). There was 100% survival in all treatments where concentrations were less than

TABLE 1. GSI and HSI (( SEM) after 18 Weeks Exposure to NP NP concentration (µg/L)a

GSI HSI

MeOH (n ) 30)

0.7 (n ) 29)

8.3 (n ) 28)

85.6 (n ) 12)

0.57 ( 0.03 1.04 ( 0.03

0.53 ( 0.03 1.07 ( 0.03

0.58 ( 0.04 0.99 ( 0.03

0.16 ( 0.02* 1.78 ( 0.05*

a Numbers of fish sampled in each treatment at this timepoint are indicated in brackets. Note that only in the fish exposed to 85.6 µgNP/L were the GSI and HSI significantly different (*, p < 0.001) from those of the controls.

8.3 µgNP/L (one fish died in the tank containing 8.3 µgNP/ L). In the tank containing the highest concentration of NP (85.6 µg/L), numbers of fish sampled were 30, 22, 19, and 12, after 0, 6, 12, and 18 weeks exposure, respectively. However, there were no significant differences in the coefficient of VOL. 35, NO. 14, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 3. Plasma vitellogenin concentrations in the fish throughout the experiment. Significant differences from the solvent control are denoted by **, p < 0.01; and ***, p < 0.001.

FIGURE 4. Plasma estradiol (E2) concentrations in the fish throughout the experiment. Significant differences from the solvent control are denoted by ***, p < 0.001. condition (k-factor) of the surviving fish maintained in the different treatments (see Figure 2). The k-factor might be expected to decrease in chronically stressed fish (34). The lack of any significant decrease in the k-factor suggests that the treated fish were not overtly affected by exposure to NP (at least with respect to somatic growth) compared to the control fish, and thus these fish were clearly not moribund. The HSI was significantly higher in the fish maintained in the highest concentration of NP than in the control fish 2912

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after 18 weeks exposure (Table 1). During the reproductive cycle of females, the HSI normally increases as vitellogenin synthesis accelerates (35, 36). The marked increase in the HSI in the fish exposed to 85.6 µgNP/L is likely to be a consequence of the large increase in vitellogenin synthesis (see below). The GSI of the fish exposed to 85.6 µgNP/L was significantly lower (p < 0.001) than that of the control fish (Table 1). In fact, ovaries of the fish exposed to 85.6 µg/L had

FIGURE 5. (a) FSHβ mRNA levels in the pituitary glands of the fish after 18 weeks of exposure to NP. Data shown represent normalized relative mRNA levels obtained by semiquantitive determination. (b) Pituitary FSH content after 18 weeks of exposure to NP. (c) Plasma FSH concentrations measured throughout the experiment. Significant differences from the solvent control are denoted by *, p < 0.05; **, p < 0.01; ***, p < 0.001. Regression analyses demonstrated a doserelated response after 6 and 12 weeks of exposure to NP. not developed at all since the start of the experiment; the GSI in the fish sampled at the onset of the experiment was 0.160 ( 0.008, and in the fish exposed to 85.6 µgNP/L for 18 weeks the GSI was 0.163 ( 0.02. There were no apparent effects of NP on ovary growth for any of the other treatment groups. There was a significant, dose-related increase in the induction of VTG in response to NP (Figure 3). The concentration of VTG in the plasma of the control fish increased, as expected, as they underwent sexual development (37). A concentration of 0.7 µgNP/L did not induce further synthesis of VTG. However, a concentration of 8.3 µgNP/L did, especially initially, when VTG concentrations were approximately 10-fold higher than the concentration in the controls. The rise in concentration of VTG in the plasma of fish exposed to the highest concentration of NP was very pronounced, and it was more than 100-fold higher than in the control fish after six weeks exposure. This set of data

FIGURE 6. (a) LHβ mRNA levels in the pituitary glands of the fish after 18 weeks of exposure to NP. Data shown represent normalized relative mRNA levels obtained by semiquantitive determination. (b) Pituitary LH content after 18 weeks of exposure to NP. (c) Plasma LH concentrations measured throughout the experiment. Significant differences from the solvent control are denoted by *, p < 0.05; **, p < 0.01; *** p. < 0.001. indicates that the NP in these tanks was behaving, at least in this respect, in an estrogenic manner. 4-Nonylphenol has been demonstrated to induce the production of VTG in vivo in male or immature female rainbow trout (16, 38). The concentration of VTG (over 100 mg/mL) in the fish exposed to the highest concentration of NP was higher than the maximum concentration found in female rainbow trout during ovarian development (which is between 50 and 100 mg/mL (37, 39)). This may be the cause of the high mortality observed in these fish, as has also been observed in juvenile female trout induced to synthesize supra-normal amounts of vitellogenin (40, 41). In those studies, excess vitellogenin (induced by exposure to exogenous E2) was found to accumulate in the liver and kidney, thus impairing the function of these organs. The experiment was initiated prior to the onset of vitellogenesis. Subsequently, the ovaries in all fish except VOL. 35, NO. 14, 2001 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

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FIGURE 7. Effects of the solvent (MeOH) on some of the parameters assessed in the study (compared with the absolute control): (a) Condition factor; (b) Plasma E2 concentration; (c) Plasma FSH concentration. Significant differences are denoted by **, p < 0.01; *** p < 0.001. those exposed to 85.6 µgNP/L increased in size as the oocytes sequestered vitellogenin. FSH has been shown to be involved with the growth of vitellogenic oocytes in fish (42), and it stimulates vitellogenic uptake into rainbow trout oocytes (43). It is possible, therefore, that the suppression in growth of the ovaries in the fish exposed to 85.6 µgNP/L can be attributed to the suppression of plasma FSH in these fish (see below). Overall, however, we can conclude that little, if any, VTG was sequestered in females exposed to 85.6 µgNP/ L, despite the very high VTG concentrations in the plasma of these fish. The expression of the vitellogenin receptor (VTGR) should have been at its highest at the stage of development at which the experiment was initiated (44), but 2914

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vitellogenin uptake was clearly inhibited in the fish exposed to 85.6 µgNP/L. Plasma E2 concentrations are shown in Figure 4. There were no differences in the concentrations of plasma E2 between groups of fish prior to dosing. Following administration of NP, fish exposed to the highest concentration had extremely low plasma E2 concentrations, which were significantly lower (p < 0.001) than those in all other groups of fish at all sampling times. Plasma E2 concentrations were not altered in the fish exposed to 8.3 µgNP/L. Plasma testosterone concentrations were measured after six weeks exposure, and were found to be suppressed in the fish exposed to 85.6 µgNP/L (data not shown). In contrast to our results, Giesy et al. (45) reported that E2 concentrations were elevated in fish exposed to NP, and they therefore suggested that NP acted indirectly as a weak estrogen. One explanation of this may be that NP leads to increased levels of aromatase, and therefore increased conversion of androgens to E2. Our finding of reduced, rather than elevated, concentrations of E2 in fish exposed to NP do not support this mechanism of action of NP; instead, they support the idea that NP acts directly on the liver through ERs, to stimulate vitellogenin synthesis. The fact that NP and other alkylphenolic chemicals can stimulate VTG synthesis in cultured hepatocytes (46) also suggests that NP acts directly. Studies on mammals also support a direct action of NP through the ER. For example, it has recently been reported that female rats dosed with OP had lowered serum E2 concentrations (47), despite showing signs of estrogenic stimulation, and a very recent paper has shown that the estrogenic activity of NP in the rat is not mediated by aromatase enzyme induction (48). The cause of the decreased plasma sex steroid titers in these fish is unknown; however, the simplest explanation is that NP causes decreased plasma FSH concentrations, which in turn prevents follicular growth. The reduced follicular growth leads to lower E2 and testosterone concentrations compared to those in the control (unexposed) fish, which increase as the follicles grow during vitellogenesis. It is also possible that NP affects the expression and/or activity of one or more of the steroidogenic enzymes; however, these parameters were not assessed in this experiment. Effects of NP on Pituitary Gland and Plasma Gonadotropins. FSH β mRNA levels in the pituitary glands of the fish are shown in Figure 5a. FSH gene expression was significantly reduced in fish exposed to NP at all the doses employed in this experiment. The amount of FSH in the pituitaries (Figure 5b) of the control fish at the termination of the experiment was 14.8 ( 1.1 µg/pituitary. This figure corresponds well with the values reported by Gomez et al. (31) for fish at a similar stage of the reproductive cycle (24 µg FSH/pituitary). A clear and significant (p < 0.001) inhibition of FSH synthesis occurred in fish maintained in the highest concentration of NP. A concentration of 8.3 µgNP/L also inhibited FSH synthesis relative to that in the control fish (p < 0.05). Plasma FSH concentrations at each timepoint in the study are presented in Figure 5c. At the last two sampling times (12 and 18 weeks), there was considerable variation in the FSH concentrations between individual fish within a group, particularly in the MeOH control fish (note the large standard errors at these times). Nonetheless, NP undoubtedly had an effect on plasma FSH concentrations; a significant (p < 0.001) inhibition of plasma FSH in the fish treated with 8.3- and 85.6 µgNP/L was observed at all timepoints; plasma FSH was also suppressed (p < 0.05) in the fish exposed to 0.7 µgNP/L at the final timepoint. These data fit well with those of the small number of studies which have demonstrated that E2 inhibits either FSH synthesis (12) or secretion (14) in rainbow trout, or FSH secretion in coho salmon (13). In our study, the response to NP of FSHβ gene expression, pituitary FSH content, and also of plasma FSH concentration (after 6 and

12 weeks of exposure) was demonstrated (by regression analysis) to be dose-related. Expression of LH mRNA in the pituitaries of the fish was suppressed in the fish exposed to 8.3 and 85.6 µgNP/L (Figure 6a). Figure 6b shows mean pituitary LH content in the fish at the end of the experiment. The mean concentration of LH in the pituitaries of the control fish was 1.79 ( 0.416 µg/ pituitary. As with the levels of FSH in the pituitary, this LH content also corresponds well with that observed for fish at this reproductive stage by Gomez et al. (31). Pituitary LH content was significantly decreased in the fish exposed to 85.6 µgNP/L. An apparent decrease in pituitary LH was also observed in the fish exposed to 8.3 µgNP/L, but this difference was not statistically significant when compared to that of the control fish. Plasma LH concentrations in the fish in our study essentially remained at baseline concentrations (0.2-0.4 ng/ mL) throughout the experiment (Figure 6c) (similar to the data presented by Prat et al. (4), and Gomez et al. (31)). The changes in plasma LH observed in our study, though statistically significant in some cases, were of small magnitude, especially when viewed in the light of the very much higher LH concentrations that occur at ovulation (around 100-fold higher than the values observed during vitellogenesis (4, 31)). It should be noted that the data obtained from this study indicated that MeOH had an effect (compared to the absolute control) on some of the parameters assessed in this study. Essentially, MeOH had no effect on growth (e.g., length, weight, or k-factor (Figure 7a)), GSI, HSI (data not shown), or pituitary gland GTH content (data not shown). However, MeOH caused a suppression of plasma E2 concentration (Figure 7b), and also an increase in plasma FSH concentration (Figure 7c); the increase in FSH may well be related to the decrease in plasma E2. These data highlight the necessity for the inclusion of solvent controls in studies such as ours, in which a relatively insoluble test chemical is added dissolved in a solvent, and hormonal endpoints are being monitored. The presence of solvent in the water, although apparently artificially challenging the fish compared to the absolute control, is not altogether unrealistic, as many effluents contain detectable concentrations of solvents (49, 50). Very few studies have investigated the influence of environmental estrogens (or other endocrine-active chemicals) on gonadotropins in fish. Khan and Thomas (21) found that o,p′-DDT (a relatively weak environmental estrogen) stimulated the release of LH in female Atlantic croaker during early recrudescence. E2 (in a separate experiment) also enhanced plasma LH concentrations. These data indicate an effect similar to that seen in our study, whereby LH release appears to be affected by an environmental estrogen. Our data also supports the work of Zilberstein et al. (18), who found a treatment of 10 µgNP/L suppressed FSHβ mRNA levels in tilapia after a three week exposure, although LHβ mRNA expression was not affected in that experiment. In contrast, pituitary LH content was enhanced in juvenile male catfish exposed to 10 µgNP/L for 14 days (19), but no effect was found in the female fish in that study. With respect to mammals, octylphenol has been shown to suppress FSH concentrations in prenatally exposed sheep (51) and in rats (52). The latter study also demonstrated a decrease of LH levels in both plasma and the pituitary gland. In summary, in this study we exposed sexually maturing female rainbow trout to a widely distributed aquatic pollutant, NP, from the stage of primary oocyte growth through to vitellogenesis. The highest concentration of this chemical employed in the study essentially shut off reproduction altogether; vitellogenin was produced in large quantities, but was not sequestered by the oocytes, which consequently did not develop further. This would clearly depress the fecundity

of these fish. The reason for this may have been the suppression of FSH synthesis (in the pituitary) and/or release (to the plasma) by NP at this concentration, since FSH is thought to induce recruitment of oocytes into the maturing pool (42), and also stimulates uptake of vitellogenin into developing oocytes (43). Lower concentrations of NP, although capable of suppressing the concentration of FSH in the plasma, did not affect the overall development of the ovary, insofar as the GSI in these fish did not differ from that of the control fish. These data demonstrate that NP is an endocrine disrupting chemical which can have multiple sites of action.

Acknowledgments C. Harris was funded by the European Commission and CEFIC. E. Santos was funded by Program PRAXIS XXI, FCT, Portugal. Thanks to Toby Carrick at the NERC Centre for Ecology and Hydrology (Windermere Laboratory) for his assistance with maintenance of the fish.

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Received for review October 30, 2000. Revised manuscript received April 10, 2001. Accepted April 16, 2001. ES0002619