ToxSci Advance Access originally published online on July 27, 2008
Toxicological Sciences 2008 106(1):93-102; doi:10.1093/toxsci/kfn150
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Environmentally Relevant Concentrations of 17
-Ethinylestradiol (EE2) Interfere With the Growth Hormone (GH)/Insulin-Like Growth Factor (IGF)-I System in Developing Bony Fish
* Division of Neuroendocrinology, Institute of Anatomy, University of Zürich, Winterthurerstr 190, CH-8057 Zürich, Switzerland
Centre de Coopération Internationale en Recherche Agronomique pour le Développement-Elevage et Médecine Vétérinaire Tropicale, Unité Propre de Recherché 20, 34398 Montpellier, France
Centre for Fish and Wildlife Health, University of Bern, Länggass-Str 122, CH-3001 Bern, Switzerland
1 To whom correspondence should be addressed. Fax: +41 44 635-5702. E-mail: reinecke{at}anatom.uzh.ch.
Received March 27, 2008; accepted July 19, 2008
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ABSTRACT |
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 TOP ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONFUNDINGREFERENCES |
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The aim of this study was to evaluate whether effects of environmental estrogens on fish growth and reproduction may be mediated via modulating the growth hormone (GH)/insulin-like growth factor I (IGF-I) system. To this end, developing male and female monosex populations of tilapia were exposed to 17
-ethinylestradiol (EE2) at 5 and 25 ng EE2/l water from 10-day postfertilization (DPF) until 100 DPF. Under exposure to both EE2 concentrations, sex ratio shifted toward more females and body length, and weight were significantly reduced in males. The growth-reducing effect was associated with significant changes in hepatic IGF-I expression, both in males and females and with significant alterations of IGF-I mRNA and GH mRNA in the brain. The changes in IGF-I and GH mRNA were accompanied by altered estrogen receptor (ER) expression in brain and liver. These findings point to an influence of estrogenic exposure on the endocrine GH/IGF-I axis. In addition, the EE2 treatment resulted in significant changes of ER and IGF-I expression in ovaries and testis, suggesting that the estrogens interact not only with the endocrine but also with the autocrine/paracrine part of the IGF-I system. Overall, our results provide evidence that EE2 at environmentally relevant concentrations is able to interfere with the GH/IGF-I system in bony fish and that the impairing effects of estrogens reported on fish growth and reproductive functions may rather result from a cross talk between the sex steroid and the IGF-I system than be toxicological. Key Words: IGF-I; growth hormone; development; liver; brain; ovary; testis; teleost.
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INTRODUCTION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONFUNDINGREFERENCES |
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Endocrine disrupting compounds (EDCs) have the potential to disrupt internal homeostasis and hormone-controlled physiological processes such as development, growth, stress responses, sexual differentiation, or reproduction. Many EDCs end up in the aquatic environment and therefore primarily affect aquatic organisms including fish. Numerous effects of EDCs on fish from all over the world have been described, such as deformities and defects in fish larvae from the North Sea (Dethlefsen et al., 1996
), feminization of male fish near sewage treatment outlets in the United Kingdom (Gross-Sorokin et al., 2006; Sumpter and Jobling, 1995), significant alterations of the steroid system of carps from the Ebro River (Lavado et al., 2004), and impaired growth in male mosquito fish in Western Australia (Game et al., 2006). Furthermore, exposure to estrogenic compounds at low concentrations can lead to growth impairment, as it has been shown for rainbow trout (Ashfield et al., 1998) and zebra fish (Schäfers et al., 2007). However, it is unknown whether the mechanisms underlying the growth impairment represents a general toxicological stress rather than an endocrine disrupting effect.
Insulin-like growth factor I (IGF-I) is crucially involved in the regulation of growth, differentiation, and reproduction by selectively promoting mitogenesis and differentiation (Jones and Clemmons, 1995; Reinecke and Collet, 1998). Among the nonmammalian classes of vertebrates, bony fish are the mostly studied with respect to IGF-I (Reinecke et al., 2005; Wood et al., 2005) mainly due to their unique development from the larval to the adult life, their high growth potential, and to their importance in aquaculture. As in mammals, IGF-I is mainly produced in fish liver––the principal source of the circulating (endocrine) IGF-I––under the influence of growth hormone (GH). In addition, IGF-I is expressed in parenchymal cells of numerous extrahepatic sites of adult (Reinecke et al., 1997) and developing (Berishvili et al., 2006a,b; Perrot et al., 1999; Radaelli et al., 2003) fish and most likely stimulates organ-specific functions by paracrine/autocrine mechanisms. There is increasing evidence that GH stimulates the expression of IGF-I not only in fish liver but also in extrahepatic sites (Biga et al., 2004; Eppler et al., 2007; Vong et al., 2003).
From in vivo and in vitro studies with adults of different fish species, initial evidence has been provided on an interaction between the estrogen and IGF-I system. In particular, 17β-estradiol (E2) appears to be able to downregulate hepatic IGF-I expression (Carnevali et al., 2005; Davis et al., 2007; Filby et al., 2006; Lerner et al., 2007; Riley et al., 2004) or serum IGF-I (Arsenault et al., 2004; Davis et al., 2007; McCormick et al., 2005). In a recent study on tilapia, Oreochromis niloticus, we could show that early life exposure to elevated concentrations of 17-ethinylestradiol (EE2), a major constituent of contraceptive pills, has long-lasting consequences on growth, IGF-I serum levels, and IGF-I expression in liver as well as in gonads (Shved et al., 2007). These data suggest that estrogen effects on the GH/IGF-I system may be more pronounced during fish development, that is, in the phase of rapid growth as well as of sexual differentiation and gonad development.
In the previous study, we used rather high EE2 concentrations (Shved et al., 2007) in order to establish principally whether the estrogen and GH/IGF-I system do interact. The effects observed with this approach, however, may reflect a general toxic stress situation rather than a specific endocrine disrupting mode of action. The aim of the present study was to investigate whether low estrogen exposure at concentrations at environmentally observed levels are able to disrupt the GH/IGF-I system and growth in developing tilapia. Thus, our hypothesis was that the EE2 impact on the GH/IGF-I system as observed at high estrogen doses is still valid at low, environmentally relevant concentrations.
It is known that EDC exposure of fish during critical stages of gonad development can induce permanent, organizational changes in sexual differentiation; however, currently very few studies have addressed whether early life exposure to EDCs, particularly estrogens, can permanently alter functioning of other physiological systems in fish (Lerner et al., 2007; Milston et al., 2003).
Concentrations of EE2 in the aquatic environment in the low nanogram per liter range have been observed in numerous countries. In Germany, concentrations of EE2 in rivers were up to 5 ng/l and in waste water treatment plants effluents ranged from 9 (Kuch and Ballschmiter, 2001) to 15 ng/l (Ternes et al., 1999). Similarly, in France and the Netherlands, 3–4 ng EE2/l were measured in coastal water and rivers and 4–8 ng EE2/l in effluents (Belfroid et al., 1999; Cargouët et al., 2004). EE2 concentrations in effluents in the United Kingdom were around 7 ng/l (Desbrow et al., 1998) and in a river in Taipei around 15 ng EE2/l (Chen et al., 2007). United States Stream Data 1999–2000 give the EE2 content in 139 streams across 30 states (Kolpin et al., 2002) between 5 and 273 ng EE2/l, and in Canada, up to 42 ng EE2/l were reported in effluents (Ternes et al., 1999).
In the present study, female and male "monosex" populations of tilapia (O. niloticus) were experimentally exposed to 5 and 25 ng EE2/l––concentrations well within the range found in the environment––from 10-day postfertilization (DPF) until 100 DPF, a period covering the sensitive period of gonadal development and ranging until the end of puberty. Male and female fish were investigated at 30, 50, 75, and 100 DPF. The question was assessed whether EE2 exposure during development exerts acute and lasting effects on the expression of GH mRNA in pituitary and of IGF-I mRNA in liver and extrahepatic sites, that is, gonad and brain. Finally, the expression of estrogen receptor (ER) that is considered to be the ER subtype most responsive to estrogen (e.g., Filby et al., 2007; Marlatt et al., 2008; Menuet et al., 2004) was also measured in the organs studied.
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MATERIALS AND METHODS |
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Generation and maintenance of fish.
To allow sex-differentiated analysis of very early stages, monosex breedings of O. niloticus generated as described before (Baroiller et al., 1995
, 1999) were investigated. In brief, suitable sires from sex-reversed fish were selected by progeny testing and then used to produce monosex populations that were used for our experiments. Fish were kept at the Aquaculture Experimental Facilities of CIRAD (Montpellier, France) where also the exposure experiments were performed.
Experimental conditions.
Two monosex populations XX and XY (n = 400 each) of tilapia were incubated in 40-l aquaria at 27°C ± 1°C during a period of 10–100 DPF with EE2 (Sigma, Deisenhofen, Germany) at the nominal concentrations of 5 or 25 ng/l and with 0.4% ETOH in water (solvent control). Each group was run in duplicates. Prior to treatment, aquaria were saturated with EE2 at a concentration of 1000 ng/l during 1 week. During the exposure, water in all aquaria (controls, 5 ng EE2/l, 25 ng EE2/l) was renewed every 48 h. The EE2 concentrations in all aquaria were repetitively determined after 36 h.
Fish sampling and tissue preparation.
At the age of 30, 50, 75, and 100 DPF fish were anesthetized with 2-phenoxy-ethanol (Sigma) added to water, and weight and length were measured. With the older stages (75 and 100 DPF), sex was additionally individually assured at sampling by external inspection of the papilla and internal inspection of the gonads. Tissue specimens (liver, brain, gonads) were excised and transferred into 1.5 ml of RNAlater (Ambion, Austin, TX), kept at 4°C and stored at – 20°C until use. Total RNA was extracted using TRIzol reagent (Invitrogen, Merelbeke, Belgium), resuspended in diethylpyrocarbonate-treated water, and photospectrometrically quantified. The following sample numbers were used for real-time PCR: control (male: 30 DPF n = 8, 50 DPF n = 12, 75 DPF n = 11, 100 DPF n = 12; female: 30 DPF n = 10, 50 DPF n = 10, 75 DPF n = 11, 100 DPF n = 12) and EE2 exposed (5 ngEE2/l, male: 30 DPF n = 8, 50 DPF n = 8, 75 DPF n = 7, 100 DPF n = 8; female: 30 DPF n = 9, 50 DPF n = 8, 75 DPF n = 10, 100 DPF n = 8; 25 ng EE2/l, male: 30 DPF n = 7, 50 DPF n = 7, 75 DPF n = 8, 100 DPF n = 8; female: 30 DPF n = 7, 50 DPF n = 8, 75 DPF n = 9, 100 DPF n = 8).
Solid-phase extraction/enzyme-linked immunosorbent assay.
To assess the actual EE2 concentration during the experiment, water samples were collected from the control and exposure aquaria. The EE2 ELISA kit (Ecologiene, Japan, purchased from Biosense Laboratories, Bergen, Norway) was used to detect EE2 in water samples. In order to concentrate the EE2 content to amounts detectable by the assay, water samples were subjected to solid-phase extraction, according to Ternes et al. (1999). In all, 100 mg LiChrolut EN and 200 mg RP-18 cartridges (Merck, Zug, Switzerland) were first conditioned with the series of passing––hexane, acetone, methanol, and water (pH 3.0). Prefiltered water samples (200 ml) were loaded under gentle vacuum pressure (1l/1h). Loaded cartridges were dried under nitrogen flow for 60 min. Elution was performed with 4 ml acetone in silanized amber vials (Supelco, Gland, Switzerland). Acetone was evaporated to dryness under nitrogen stream, and EE2 was reconstituted in 0.5 ml ethanol. Resulting samples were analyzed by enzyme-linked immunosorbent assay.
Design of primers and probes for real-time PCR.
Based on the mRNA sequences of O. niloticus β-actin as housekeeping gene (Hwang et al., 2003), O. mossambicus IGF-I (Reinecke et al., 1997), O. niloticus GH (Ber and Daniel, 1992), and O. niloticus ER (Chang et al., 1999), primers (β-actin: sense GCCCCACCTGAGCGTAAATA, antisense AAAGGTGGACAGGAGGCCA, probe TCCGTCTGGATCGGAGGCTTCATC; IGF-I: sense TCTGTGGAGAGCGAGGCTTT, antisense CACGTGACCGCCTTGCA, probe ATTTCAATAAACCAACAGGCTATGGCCCCA; GH: sense TCGACAAACACGAGACGCA, antisense CCCAGGACTCAACCAGTCCA, probe CGCAGCTCGGTCCTGAAGCTG, and ER: sense CAAGTGGTGGAGGAGGAAGATC, antisense CTCAGCACCCTGGAGCAG, probe CTGATCAGGTGCTCCTC) were created as described previously (Caelers et al., 2004; Shved et al., 2007). Oligonucleotides were designed with Primer Express software version 1.5 (PE Biosystems, Foster City, CA).
Quantitation of IGF-I and ER expression by two-step real-time reverse transcriptase-PCR TaqMan system.
Total RNA was treated with 1 U of RQ1 RNAse-free DNAse (Catalys AG, Wallisellen, Switzerland). Single-stranded cDNA was synthesized from 800 ng total RNA using 1x TaqMan reverse transcription buffer, MgCl2 (5.5mM), 1.25 U/µl of moloney murine leukemia virus reverse transcriptase, 2.5µM of random hexamers primers, 0.4 U/µl ribonuclease inhibitor, 500µM each dNTP (Applied Biosystems, Rotkreuz, Switzerland) for 10 min at 25°C, 30 min at 48°C, and 5 min at 95°C. In all, 2 µl cDNA obtained from 10 ng/µl total RNA were subjected, in duplicates, to real-time PCR using an Absolute QPCR low ROX Mix (ABgene, Epsom, UK) including Thermo-Start DNA polymerase, 300nM of each primer, and 150nM of the fluorogenic probe. Amplification was performed with a total reaction volume of 10 µl in a MicroAmp Fast Optical 96-Well Reaction Plate (Applied Biosystems). Reactions were run on the ABI 7500 Fast Real-Time PCR System (Applied Biosystems) with the following conditions: 95°C for 15 min of the enzyme activation followed by 40 cycles of 95°C for 15 s and 60°C for 1 min.
Relative quantification of treatment effects using the 
CT method.
The relative gene expression ratios between EE2-exposed and control fish were calculated using the comparative CT method (CT method). For validation of quantitative PCR assays, the CT values were plotted against the logarithms of the dilution factors and the slope was determined. Relative changes induced by EE2 exposure were calculated using the formula 
with CT = CT (target gene) – CT (internal control) and CT = CT (treated group) – CT (untreated control). Thus, all experimental data are expressed as an n-fold difference relative to the calibrator (untreated control). Statistical significance was calculated using Mann-Whitney rank sum test with an exact p value. Statistical analyses were performed with GraphPad Prism 4.
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RESULTS |
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Actual EE2 Concentrations in the Experimental Tanks
Measurements of the mean actual EE2 concentrations revealed in controls 0.009 ± 0.013 ng EE2/l, in the 5-ng EE2/l aquaria 4.34 ± 0.85 ng EE2/l, and in the 25-ng EE2/l aquaria 16.23 ± 5.74 ng EE2/l.
Sex Ratio
In our study, we have used so-called monosex populations of tilapia. In untreated controls, the percentage of females in the female monosex group amounted to about 76% and that of males in the male monosex group to 80%. A population with 75% or more fish of the same sex is considered to be monosex (D'Cotta et al., 2001). Thus, our percentage is well within the range that can be achieved in tilapia (Baroiller et al., 1995) because sex chromosome–linked genetical factors determine sex largely but not exclusively, and the development of final sex phenotype is influenced by autosomal genetic factors in combination with the hormonal, behavioral, and environmental cues (Baroiller et al., 1999; Devlin and Nagahama 2002).
Both, the female and male monosex fish populations exhibited a higher percentage of females after exposure to both 5 and 25 ng EE2/l, the changes being significant in female monosex (Table 1).
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Body Length and Weight
Exposure to EE2 affected growth more pronouncedly in male than in female fish. In male fish (Fig. 1a–c), from 50 DPF on (i.e., 40 days after start of exposure), both body length and weight were significantly reduced in fish exposed either to 5 or 25 ng EE2/l. At the end of the experiment (100 DPF), body length of males (Fig. 1a) was reduced by about 9.5 (5 ng EE2/l) and 11% (25 ng EE2/l) and body weight (Fig. 1c) by about 25% at both EE2 concentrations (Fig. 1). In female fish (Figs. 1d–f), both parameters showed only a trend to decrease. Here at 100 DPF, body length (Fig. 1d) was reduced by about 3 (5 ng EE2/l) and 6.2% (25 ng EE2/l) and body weight (Fig. 1f) by about 14% at 25 ng EE2/l.
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Liver IGF-I and ER
Gene ExpressionIn both sexes, an initial (30 DPF) significant upregulation of IGF-I mRNA levels (5 ng EE2/l, male: 2.82-fold, p = 0.0004; female: 2.43-fold, p = 0.003; 25 ng EE2/l, male: 1.74-fold, p = 0.05, female: 2.16-fold, p = 0.006) under both EE2 concentrations and in both sexes was observed (Figs. 2a and 2c). This was accompanied by elevated hepatic ER
levels (Figs. 2b and 2d) in both males and females at 5 ng EE2/l (male: 1.77-fold, p = 0.003; female: 2.24-fold, p = 0.0007) and in females (Fig. 2d) also at 25 ng EE2/l (2.66-fold, p = 0.0007).
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During prolonged exposure (i.e., at 50, 75, and 100 DPF), the initial EE2-related upregulation of liver IGF-I mRNA changed into a downregulation. In males, IGF-I gene expression was significantly reduced under 25 ng EE2/l (– 2.38-fold, p = 0.0003) at 50 DPF (Fig. 2a) and with both EE2 concentrations at all subsequent time points. Changes in IGF-I gene expression were paralleled by significant downregulations of ER
gene expression levels (Fig. 2b) at 75 DPF (5 ng EE2/l: – 1.64-fold, p = 0.0047; 25 ng EE2/l: – 2.7-fold, p = 0.0003) and 100 DPF (5 ng EE2/l: – 2.08-fold, p = 0.0082).
Similar dynamics like in males, although delayed, were observed in females: suppressed IGF-I mRNA levels (Fig. 2c) occurred in females at 75 (25 ng EE2/l: – 1.41-fold, p = 0.0093) and 100 DPF (5 ng EE2/l: – 1.85-fold, p = 0.015). The observed lowered IGF-I mRNA levels, in part, coincided with suppression (100 DPF, 25 ng EE2/l: – 1.59-fold, p = 0.05) of ER mRNA (Fig. 2d).
Brain IGF-I and ER Gene Expression
In male brains, IGF-I mRNA levels were significantly suppressed at 30 (– 1.41-fold, p = 0.0499) and 50 DPF (– 1.72-fold, p = 0.0148) at the higher EE2 concentration (25 ng EE2/l) but returned to control levels at 75 DPF (Fig. 3a). ER mRNA was significantly elevated at 30 DPF at both concentrations (5 ng EE2/l: 1.76-fold, p = 0.0499; 25 ng EE2/l: 1.73-fold, p = 0.0379) and at 75 DPF at 5 ng EE2/l (1.65-fold, p = 0.0205) (Fig. 3b). At end of the experimental period (100 DPF), a significant downregulation (– 1.89-fold, p = 0.0011) of ER mRNA was observed under 5 ng EE2/l (Fig. 3b).
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In female brains, IGF-I gene expression levels (Fig. 3c) revealed an initial upregulation at 30 DPF (25 ng EE2/l: 2.16-fold, p = 0.0012) followed by significant suppression at 50 DPF at both concentrations (5 ng EE2/l: – 1.59-fold, p = 0.0002; 25 ng EE2/l: – 1.47-fold, p = 0.0133). Under 5 ng EE2/l, IGF-I mRNA was still reduced at 75 DPF (– 2.38-fold, p = 0.0001) followed by a significant elevation at 100 DPF (1.66-fold, p = 0.0281). The ER
mRNA levels tended to be suppressed over the whole treatment phase. At several points in time (5 ng EE2/l at 50 DPF: – 1.53, p = 0.0152; 25 ng EE2/l at 100 DPF: – 1.41-fold, p = 0.0496), the suppression was significant (Fig. 3d).
Gonad IGF-I and ER Gene Expression
In male gonads, IGF-I mRNA levels (Fig. 4a) were significantly reduced at 50 and 75 DPF (50 DPF: 5 ng EE2/l: – 3.04-fold, p = 0.0286; 25 ng EE2/l: – 6.07-fold, p = 0.0286; 75 DPF: 25 ng EE2/l: – 2.33-fold, p = 0.0303). Only at 100 DPF, IGF-I mRNA levels returned to control levels (5 ng/l) or were even significantly elevated (25 ng/l). ER mRNA in testes was suppressed over the whole exposure period. This decrease was statistically significant at 30 DPF under 25 ng EE2/l (– 2.09-fold, p = 0.0339) and at 50 (– 2.73-fold, p = 0.0286) and 100 DPF (– 1.95-fold, p = 0.037) under 5 ng EE2/l (Fig. 4b).
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In female gonads, IGF-I mRNA expression was significantly elevated at 100 DPF (5 ng/l: 3.17-fold, p = 0.0023; 25 ng EE2/l: 2.55-fold, p = 0.0042) (Fig. 4c). ER
mRNA levels did not change throughout the experimental period (Fig. 4d).
Pituitary GH Levels
The amount of GH mRNA was determined using extracts from whole brain with pituitary attached since the pituitaries at 30 and 50 DBF were too small to allow dissection. At 30 DPF, no effect of EE2 exposure on GH expression was observed in both sexes (Fig. 5). In females, after a significant increase in GH mRNA at 50 DPF (5 ng EE2/l: 3.91-fold, p = 0.002; 25 ng EE2/l: 2.62-fold, p = 0.05), GH mRNA decreased at 100 DPF under 5 ng EE2/l to the threefold (p = 0.03) and under 25 ng EE2/l to the 10-fold (p = 0.03) (Fig. 5a). In males, GH mRNA exhibited only a trend to similarly increase at 50 DPF but was significantly decreased at 100 DPF under 25 ng EE2/l by the sevenfold (p = 0.001) (Fig. 5b).
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DISCUSSION |
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TOPABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTSDISCUSSIONFUNDINGREFERENCES |
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Developmental exposure to EE2 clearly shifted the sex ratio: female and male monosex fish populations exhibited a higher percentage of females after exposure to both 5 and 25 ng EE2/l. The growth rate of the developing tilapia was also suppressed by EE2 treatment. These findings are in agreement with our previous observations on high-dose effects of EE2 (Shved et al., 2007
), but the present results indicate that concentrations at environmental levels of EE2 still have the capacity to impair tilapia growth and sexual development.
Our hypothesis was that the growth reduction resulting from estrogen exposure is related to a cross talk between the estrogen and the IGF-I system, be it through an estrogen effect on the liver as source of the circulating (endocrine) IGF-I, or be it an estrogen effect on the paracrine/autocrine IGF-I system in the peripheral organs. Our experimental results show that hepatic IGF-I expression can be suppressed by EE2 exposure. This is particularly the case in males but with prolonged exposure also in females. The decline in hepatic IGF-I mRNA found here in developing tilapia is in agreement with previous investigations in which estrogen exposure of adults of different fish species led to a decrease in hepatic IGF-I mRNA (Carnevali et al., 2005; Filby et al., 2006; Riley et al., 2004). The stronger effect on liver IGF-I of males than of females correlates with a stronger EE2 effect on growth indices in males. This correlation suggests that the EE2-induced growth reduction goes back to the EE2-induced reduction of hepatic IGF-I expression, that is, to an endocrine effect and not to a general toxic effect.
This interpretation is further supported by our findings on GH mRNA, which reacts roughly in parallel to the endocrine hepatic IGF-I system. In females, GH mRNA was significantly raised at 50 DPF but not in males. The different response also may explain the lesser growth impairment found in females when compared to males. At 100 DPF, in both sexes, GH mRNA was significantly downregulated. This decrease in GH mRNA may well be caused by a direct action of EE2 on the pituitary GH cells. Because E2 increased the expression of somatostatin-14 in goldfish brain (Canosa et al., 2002) EE2 may have as well suppressed GH at the hypothalamic level by acting on somatostatin-14.
Very few studies have dealt with effects of estrogens on fish pituitary GH. No effects on GH expression were observed in adult fathead minnow exposed to 35 ng E2/l in surrounding water for 14 days (Filby et al., 2006). However, in steroid-primed immature rainbow trout after E2 injections (Holloway and Leatherland, 1997) and in female goldfish after implantation of pellets containing 100 µg E2/g body weight for 5 days (Zou et al., 1997), a stimulation of GH synthesis and secretion by E2 has been reported. Similarly, in vitro treatment of rainbow trout pituitary glands with 1000nM of the xenoestrogen o,p'-dichlorodiphenyltrichloroethane resulted in a significant induction of GH mRNA (Elango et al., 2006). The results of the latter studies reporting an increase in GH mRNA after short-term treatment with (xeno)estrogens are consistent with ours that show the significant increase in GH mRNA in females at 50 DPF and a similar trend in males. While Shved et al. (2007) did not investigate early stages, they reported that feeding of tilapia with 125 µg EE2/g food from 10 to 40 DPF led to a significant reduction of pituitary GH mRNA female (75 and 90 DPF) and male (165 DPF) fish. This reduction is compatible with the present results where a significant decrease in GH mRNA occurred after 90 days of EE2 exposure. Some remaining discrepancies between above studies and the present report are probably due to the different estrogenic compounds used, to the different modes of application, or to hormone application at different life stages of the fish.
The question arises as to whether the altered hepatic IGF-I expression in EE2-exposed fish is a direct consequence of the EE2-mediated activation of the ER pathway. To this end, we analyzed the expression of ER. In our study, estrogen treatment led to a downregulation of hepatic ER expression what is in contrast to the studies of Filby et al. (2006) and Shved et al. (2007). In both these studies, estrogen-induced decrease in IGF-I mRNA was accompanied by an upregulation of ER mRNA. However, a major difference between our study and these two is that we used a long-term exposure regime, while those treated the fish only for 14 and 30 days, respectively. At 30 DPF, that is, after 20 days of exposure, we also observed an upregulation of hepatic ER, and only thereafter, it turned into a downregulation. Thus, the difference between our results and those of Filby et al. (2006) and Shved et al. (2007) seems to be due to the exposure duration. The hepatic IGF-I levels showed time-dependent changes that are fairly similar to those of the ER, with an initial upregulation followed by a downregulation on the longer run. This correlation suggests a mechanistic relationship between the ER pathway and the IGF-I gene expression, but future studies will have to substantiate this hypothesis.
In contrast to the liver, the peripheral organs show quite different responses of ER and IGF-I mRNA to EE2 treatment suggesting that the mechanisms of the IGF-I-estrogen cross talk differ between the endocrine part of the IGF-I system (liver) and the autocrine/paracrine part (gonads and brain). ER mRNA in male brain was significantly elevated at 30 DPF at both concentrations to be significantly decreased at the end of the experimental period (100 DPF) under 5 ng EE2/l. In females, brain ER mRNA tended to be decreased with some significancies all over the experimental period. In contrast, in adult fathead minnow, exposure to 35 ng E2/l for 14 days did not influence ER mRNA in brain of either sex (Filby et al., 2006). Thus, the brain ER mRNA seems to be downregulated only during long-term developmental exposure. In the same study (Filby et al., 2006), IGF-I mRNA in brain was significantly induced in both sexes, the increase being more pronounced in female brain. This is quite consistent with the present data where also a significant upregulation of IGF-I mRNA in female brain was observed after 20 days of EE2 exposure. Later, however, brain IGF-I mRNA was downregulated, the effect being stronger in females.
Developmental exposure to low concentrations of EE2 has also consequences on the IGF-I expression in organs where this hormone is considered to have an autocrine or paracrine function such as brain and gonads. The response patterns in these organs were different to the liver. However, also here a sex-specific response pattern was indicated, at least in the brain as outlined above. In the testis, IGF-I showed a significant but transitory downregulation which disappeared or even turned into an induction under prolonged exposure. In contrast, in the ovaries of exposed fish, IGF-I mRNA levels showed an increase becoming significant at 100 DPF. The functions of IGF-I include roles in spermatogenesis and oocyte proliferation and maturation. In Japanese eel and tilapia cultured testes, IGF-I stimulated spermatogenesis induced by 11-ketosterone (Nader et al., 1999; Tokalov and Gutzeit, 2005). In rainbow trout testes, IGF-I mRNA increased after GH treatment (Biga et al., 2004; Le Gac et al., 1996) and GH and IGF-I stimulated the incorporation of thymidine into spermatogonia and primary spermatocytes (Loir, 1999). In the ovary of different fish species, IGF-I stimulated thymidine incorporation in vitellogenic follicles (Srivastava and Van der Kraak, 1994), promoted oocyte maturation (e.g., Lokman et al., 2007; Negatu et al., 1998), and selectively influenced the production of sex steroids (e.g., Chourasia and Joy, 2007; Weber et al., 2007). In sturgeon, previtellogenic follicles exhibited higher IGF-I and IGF-I receptor (IGF-1R) mRNA levels in maturing females than in age-matched non-maturing females, and females entering vitellogenesis showed an increase in ovarian IGF-I and IGF-1R mRNA (Wuertz et al., 2007), thus emphasizing the importance of local IGF-I. Recent evidence obtained in zebra fish exposed to 10 ng EE2/l surrounding water indicated not only a reduction of juvenile growth but also an impairment of sexual maturity, adult fecundity, and fertilization success (Schäfers et al., 2007). Similarly, the exposure of Japanese medaka to environmentally relevant concentration of 1.6–157 ng/l E2 impaired fecundity and reduced the number of spawns (Jukosky et al., 2008). Whether the EE2 effect on IGF-I expression in gonads of developing tilapia is involved in the feminization effect of the experimental treatment cannot be asserted from the present data; however, it is clear that EE2 changes normal expression patterns of IGF-I in gonad ontogeny and thereby has the potential to lead to adverse consequences.
Our study advances endocrine disruptor research in that it extends the targets for action of EDCs beyond the reproductive axis––which has received most attention to date. Importantly, we could demonstrate that such effects occur at low concentrations similar to environmental levels. Individual growth has consequences for demographic parameters such as age-specific survival, time to maturation, or fecundity. Thus, our findings obtained with low doses of EE2 point to the impact of the GH/IGF-I system, through which environmental estrogens, in addition to their direct effect on fish reproduction, could alter population growth of fish species. Future research has to address the mechanisms through which estrogens act to modulate the GH/IGF-I system. In particular, given our findings on the correlation between ER and IGF-I changes, it would be important to examine whether the EE2 effects on the GH/IGF-I system in liver and extrahepatic sites are mediated through the ER pathway and/or through other mechanisms.
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The Swiss National Research Foundation (NRP 50, project 4050-66580).
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