Neural aromatase transcript and protein levels in Atlantic salmon (Salmo salar) are modulated by the ubiquitous water pollutant, 4-nonylphenol

General and Comparative Endocrinology 164 (2009) 91–99

Contents lists available at ScienceDirect

General and Comparative Endocrinology journal homepage: www.elsevier.com/locate/ygcen

Neural aromatase transcript and protein levels in Atlantic salmon (Salmo salar) are modulated by the ubiquitous water pollutant, 4-nonylphenol Trond M. Kortner *, Anne S. Mortensen, Marianne D. Hansen, Augustine Arukwe Department of Biology, Norwegian University of Science and Technology (NTNU), Høgskoleringen 5, 7491 Trondheim, Norway

a r t i c l e

i n f o

Article history: Received 30 March 2009 Revised 11 May 2009 Accepted 19 May 2009 Available online 23 May 2009 Keywords: Aromatase Brain Endocrine-disrupting chemicals Xenoestrogens

a b s t r a c t At present, there are no known direct occurrences of nonylphenol (NP) in nature. Therefore, its presence in nature is solely a consequence of human activities. NP is generated through degradation of alkylphenol ethoxylates released mainly from textile, metal working, institutional cleansing and laundry cleaning, but few data on the amount of the release is available. These compounds have been shown to affect several biological processes, including the endocrine systems, in a wide number of species. The cytochrome P450 aromatase (Cyp19) is the rate-limiting step in estrogen production, and is known to be a potential target for endocrine-disrupting chemicals (EDCs) such as NP. Teleost fish generally have a high brain aromatase activity, and the effects of EDCs in fish brain is not thoroughly investigated. In this study, juvenile Atlantic salmon (Salmo salar) were exposed to waterborne concentrations of the synthetic pharmaceutical and xenoestrogen 17a-ethynylestradiol (EE2; 5 ng/L) and the xenoestrogen 4-nonylphenol (NP; 5 and 50 lg/L) for 72 h. Brain tissue and blood were sampled from individual fish. Gene expression patterns of Cyp19 isoforms were determined by quantitative PCR, aromatase protein immunoreactivity in the brain was evaluated by immunohistochemistry and immunoblotting, and aromatase activity was analyzed using the tritiated water-release assay. Plasma estradiol (E2) and testosterone (T) levels were measured by EIA. In the brain, EE2 increased the mRNA expression of Cyp19b almost threefold compared to the solvent control, whereas Cyp19a levels were unaffected by EE2 treatment. In contrast, both NP concentrations produced significant reduction of Cyp19a expression. Immunohistochemical aromatase protein reactivity was localized in several brain regions, but no apparent quantitative effects of the exposures were observed. Immunoblotting analysis showed that EE2 and NP produced a slight increase in brain immunoreactive aromatase protein band, compared with controls. Plasma levels of E2 increased twofold when treated with EE2 and 5 lg NP/L, and threefold when exposed to 50 lg NP/L. In general, the present study shows that the parallel biochemical, transcriptional and cellular detection of neural aromatase for endocrine-disrupting effects from EE2 and NP may be observed at specific levels of the biological organization. Ó 2009 Elsevier Inc. All rights reserved.

1. Introduction Certain contaminants in the environment act in similar ways as hormones, with resulting abilities of influencing endocrine control and regulation of integral biological functions. A large number of these contaminants are produced for specific purposes and are used in pesticides, plastics, surfactants, electrical transformers and other products (for review see Ahlborg et al., 1995; Caldwell, 1986). Other substances are generated as by-products during manufacturing or are breakdown products of some other chemical (e.g., nonylphenol, NP), and some, such as 17b-ethinylestradiol (EE2) and diethylstilbestrol (DES), are drugs. Some estrogenic contaminants are natural compounds such as the phytoestrogens and mycoestrogens that occur in a variety of plants and fungi (Arukwe * Corresponding author. Fax: +47 73 591309. E-mail address: [email protected] (T.M. Kortner). 0016-6480/$ - see front matter Ó 2009 Elsevier Inc. All rights reserved. doi:10.1016/j.ygcen.2009.05.009

et al., 1999; Celius et al., 1999; Pelissero et al., 1991, 1993). Regardless of the source or original intended use, substantial amounts of these chemicals end up in the marine and freshwater environments due to physico-chemical, hydrologic and atmospheric processes (Guardans and Gimeno, 1994). For example, it has been established that alkylphenol polyethoxylates (APEs) undergo a complex microbial degradation pattern, characterized by the formation of several metabolic products, including NP, that are more toxic than the parent compound (Ekelund et al., 1990; Naylor et al., 1992). NP is used as raw materials, for surface-active agents (anionic surface active agent and nonionic surface active agent), ethyl cellulose stabilizers, oil soluble phenyl resins and esters. It is also used as processed articles, for detergents, oil varnishes, rubber auxiliaries and vulcanization accelerators, antioxidants and corrosion inhibitors for petroleum products, sludge generation inhibitors for petroleum. Thus, NP and APEs are ubiquitous marine and freshwater contaminants that represent serious environmental

92

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

health concerns (Ahel et al., 1994). Because of its enhanced resistance toward further biodegradation, toxicity, estrogenic effects and ability to bioaccumulate in aquatic organisms, NP have been identified as the most critical metabolite of APEs in several studies (Ahel et al., 1994). Levels of alkylphenolic compounds as high as 180 lg/L have been detected in rivers in the United Kingdom (UK; Blackburn and Waldock, 1995) 60 lg/L in rivers in Switzerland (Ahel et al., 1994) and 17 lg/L in a river estuary in Croatia (Kvestak et al., 1994). Since high concentrations of alkylphenolic compounds are found in estuaries, it is plausible to assume that these compounds could impose an estrogenic risk to the aquatic organisms, a risk that may not be restricted to areas close to the outlets of municipal and industrial effluence. EE2 is a pharmaceutical and potent endocrine modulator known to be present in the aquatic environment at biologically active concentrations (Nash et al., 2004; Rotchell and Ostrander, 2003). The concentration of EE2 reported in effluents and surface waters from Europe range between 0.5 and 7 ng/L (Larsson et al., 1999) and concentrations of up to 50 ng/L have been reported (Aherne and Briggs, 1989). In the United States, a survey of 139 streams showed that several rivers had concentrations >5 ng/L with an extreme EE2 concentration up to 273 ng/L reported at some riverine sites (Kolpin et al., 2002). Despite the lower EE2 concentrations in surface waters compared to natural steroidal estrogens, its estrogenic potency in fish in vivo studies is 10- to 50-fold higher than that of estradiol-17b (E2) and estrone (E1) (Segner et al., 2003) most likely due to its longer half-life and tendency to bioconcentrate (650- and 10,000-fold in whole-body tissues and bile, respectively) (Larsson et al., 1999). The cytochrome P450 aromatase (Cyp19) is a crucial enzyme in the steroidogenic pathway that catalyzes the conversion of androgens to estrogens. As Cyp19 catalyzes the rate-limiting step in estrogen production, it is the main regulator of local and systemic estrogen levels in the body, and plays integral roles in reproduction, sexual differentiation and male/female behavior (Cheshenko et al., 2008). Additionally, androgens and estrogens seem to have general metabolic roles that are not directly involved in reproductive processes (Simpson et al., 2002). Although the gonads are considered to be the main site for estrogen production, teleosts are known for exhibiting unusually high neural Cyp19 mRNA and activity levels (Forlano et al., 2001; Pasmanik and Callard, 1985). In contrast to mammalian and avian species, most teleost species have two different Cyp19 genes that encode two functionally distinct aromatase isoforms, namely – Cyp19a and Cyp19b. Cyp19a is the dominant ovarian isoform, whereas Cyp19b is mostly expressed in brain (Callard et al., 2001). In most teleost species, while brain Cyp19 activity is highest in the forebrain, including the telencephalon, preoptic area and hypothalamus (Pellegrini et al., 2005), Cyp19b activity and mRNA seem restricted to radial glial cells (Forlano et al., 2001; Menuet et al., 2003). Furthermore, brain Cyp19 activity and mRNA levels fluctuate seasonally and with reproductive state, and seem to mirror changes in circulating steroids (Forlano et al., 2006). The regulation of Cyp19 mRNA expression and activity is not limited to naturally occurring steroids and is known to be a potential target for numerous endocrine-disrupting chemicals (EDCs) (Kishida et al., 2001; Menuet et al., 2005). Previously, we showed that the synthetic pharmaceutical (EE2) and the xenoestrogen NP modulated salmon brain Cyp19 expression (Lyssimachou et al., 2006; Meucci and Arukwe, 2006a) and that a substantial amount of NP accumulated in the central nervous system of juvenile salmon (Arukwe et al., 2000). But, the cellular localization and the differences in expression patterns in the brain after exposure to xenoestrogens in relation to mRNA, protein and activity levels have not been demonstrated in teleost species. Therefore, the present study was undertaken in order to establish whether the previously

observed xenoestrogen-induced alterations in brain Cyp19 and other steroidogenic genes will parallel modulation of functional protein levels, activity or designate to particular cellular localization in the brain. Furthermore, we wanted to evaluate whether these modulations will parallel alterations in plasma steroid hormone levels. Since it has been well established that neurosteroid hormones have multiple functional and structural effects during development and adulthood (Baulieu et al., 2001; Schumacher et al., 1996), the cellular detection of aromatase expression may provide sensitive tools for evaluating water quality with added biological significance. 2. Materials and methods 2.1. Chemicals and reagents 4-Nonylphenol (85% of p-isomers) was purchased from Fluka chemika-biochemika (Buchs, Switzerland). 17a-Ethynylestradiol, Trizol reagent and TA cloning kit were purchased from Gibco-Invitrogen life technologies (Carlsbad, CA, USA). 3H-andreostenedione bovine serum albumin (BSA) was purchased from Sigma–Aldrich Chemical Company (St. Louis, MO, USA). Iscript cDNA Synthesis Kit and iTaqTMSybrÒ Green supermix with ROX were purchased from Bio-Rad laboratories (Hercules, CA, USA). SuperPicTureTM polymer detection kit was purchased from Zymed (San Francisco, CA, USA). Testosterone (T) and estradiol-17b (E2) enzyme immunoassay (EIA) kits were purchased from Cayman chemical company (Ann Arbor, MI, USA). All other chemicals were of the highest commercially available grade. 2.2. Animals and treatments Immature Atlantic salmon (mean weight and length 9.2 ± 1.3 g and 9.4 ± 2.4 cm, respectively) were obtained from Lundamo hatcheries (Trondheim, Norway) and kept in 70-L aquariums at 7.0 ± 0.5 °C with a 12:12-h photoperiod at the Department of Biology, Norwegian University of Science and Technology (NTNU) animal-holding facilities. A 5-day acclimation period proceeded the exposure period. Fish were exposed to waterborne concentrations of EE2 (5 ng/L) and NP (5 and 50 lg/L). Control fish were only exposed to the solvent (ethanol, 0.001%). After 72 h acute exposure, fish were anesthetized with isoeugenol (25 mg/L), weighed, and blood samples were taken from the caudal vein using heparinised syringes. Blood samples were centrifuged and obtained plasma was stored at 80 °C until further analyses. Thereafter, fish was sacrificed and the brain and liver were rapidly excised and weighed. The brain tissues were sectioned into its left and right hemispheres. The right hemisphere was fixed in 4% paraformaldehyde and used for histological analyses, whereas the left hemisphere was homogenized in Trizol reagent, snap frozen in liquid nitrogen and stored at 80 °C until further processing. Additional fish brains for aromatase activity assays and immunoblotting were sampled, snap frozen in liquid nitrogen and stored at 80 °C until further processing. Liver tissue was cut into suitable pieces, homogenized in Trizol reagent, snap frozen in liquid nitrogen and stored at 80 °C until further processing. In all analysis, the number of samples was n = 5, unless otherwise stated. 2.3. Quantitative (real-time) PCR Total cDNA for the real-time PCR reactions were generated from 1 lg total RNA from all samples using a combination of poly-T and random primers from an iScript cDNA Synthesis Kit as described by the manufacturer (Bio-Rad). High quality RNA with A260/A280 ratio above 1.9 and intact ribosomal 28S and 18S RNA bands (evaluated by formaldehyde agarose gel electrophoresis) was used for

93

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

cDNA synthesis. PCR primers for amplification of 96–391 bp genespecific PCR-products were designed from conserved regions of the studied genes. The primer sequences, their amplicon size and the optimal annealing temperatures are shown in Table 1. Prior to PCRs, all primer pairs were used in titration reactions in order to determine optimal primer pair concentrations and their optimal annealing temperatures. All chosen primer pair concentrations used at the selected annealing temperatures gave a single band pattern for the expected amplicon size in all reactions. PCR products from the genes to be investigated were cloned into pCR2.1 vector in INVaF’ Escherichia coli (Invitrogen). Each plasmid was sequenced by Eurofins MWG Operon (Ebersberg, Germany). All sequences were confirmed using NCBI nucleotide BLAST software. Quantitative (real-time) PCR was used for evaluating gene expression profiles. For each treatment, the expression of individual gene targets was analyzed using the Mx3000P Real-Time PCR System (Stratagene, La Jolla, CA, USA). Each 25-lL DNA amplification reaction contained 12.5 lL of iTAQTMSYBR Green Supermix with ROX (Bio-Rad), 1 lL of cDNA, and 200 nM of each forward and reverse primer. The three-step real-time PCR program included an enzyme activation step at 95 °C (3 min) and 40 cycles of 95 °C (15 s), 55–60 °C (30 s) (depending on the primers used; see Table 1), and 72 °C (20 s), and a dissociation curve was included in the assays. Controls lacking a cDNA template were included in all qPCR analysis as a quality control. Cycle threshold (Ct) values obtained were converted into mRNA copy number using standard plots of Ct versus log copy number. The criterion for using the standard curve is based on equal amplification efficiency (usually >90%) with unknown samples and this is checked prior to extrapolating unknown samples to the standard curve. The standard plots were generated for each target sequence using known amounts of plasmid containing the amplicon of interest, as described previously (Arukwe, 2005). Data obtained from triplicate runs for target cDNA amplification were averaged and expressed as percentage of control samples. 2.4. Antibody production and immunoassays A polyclonal aromatase antibody was produced by immunization of rabbits with a synthetic peptide that was produced based on a conserved region of the cyp19 gene from several species (Genscript, Piscataway, NJ, USA). The peptide sequence was: H2N–KPN EFS LEN FDK TVC–CONH2 with a molecular weight of 1.771 kilodalton (kDa). Rabbits were immunized once a week with the synthetic peptides and were bled after the fourth boost. The resulting sera were purified using Hitrap rProtein A affinity column (Amersham, Uppsala, Sweden). Brain tissues were fixed in 4% paraformaldehyde in 0.1 M sodium phosphate buffer (pH 7.4) for 24 h. After fixation, tissues were washed twice in 70% ethanol, dehydrated in a graded series of ethanol baths using an automatic tissue processor (Leica Microsystems, Nussloch, Germany), embedded in paraffin blocks, and sagittal sections (4 lm) were stained with Mayer’s hematoxylin and eosin for routine histology. Immunohistochemical analysis was performed on sagittal sections (4 lm) as described previously (Kortner and Arukwe, 2007) using an overnight incubation at room temperature of a 1:100 dilution using a polyclonal antibody as described above.

To verify the specificity of the anti-aromatase antiserum used for immunohistochemistry, Western blotting was performed. Tissues for immunochemical studies were homogenized in 4 volumes 0.1 M sodium phosphate buffer (containing 0.15 M KCl, 1 mM EDTA, 1 mM dithiothreitol (DTT) and 10% glycerol at pH 7.4). The homogenate was centrifuged for 20 min at 12,000 g at 4 °C, and the pellets were discarded. Total protein concentrations in samples were determined with the Bradford method (Bradford, 1976) using bovine serum albumin (BSA) as standard. Proteins (10 lg) were separated by 9% sodium dodecyl sulfate polyacrylamide gel electrophoresis (SDS–PAGE). The gel was blotted onto polyvinylidene difluoride (PVDF) membranes (Bio-Rad) and incubated with the primary aromatase antiserum diluted 1:250. After washing, membranes were incubated with peroxidase conjugated goat anti-rabbit antibodies (GAR-HRP; Bio-Rad) diluted 1:3000, developed using an Immun-Star WesternC Chemiluminescent Kit (Bio-Rad) and visualized with Eastman KODAK Company’s Molecular Imaging Systems (Rochester, NY, USA). A quantitative vitellogenin (Vtg) ELISA was performed as previously described (Mortensen and Arukwe, 2007), to demonstrate the estrogenicity of EE2 and NP. 2.5. Aromatase activity assay Aromatase activity was determined using the tritiated waterrelease assay as previously described (Lyssimachou et al., 2006). Briefly, brain tissue from individual fish (20–40 mg) was homogenized in 100 lL of 10 mM phosphate buffer (containing 100 mM EDTA, 1 mM dithiothreitol at pH 7.4) per milligram of tissue and centrifuged at 10,000g for 20 min at 4 °C. A 50-lL volume of the supernatant was incubated in a water bath at 25 °C for 1 h together with 4 nM of 3H-androstenedione and 1 mM NADPH in a total volume of 200 ll. The samples were assayed in duplicates with a blank containing buffer instead of sample for each set. The reaction was stopped by adding 300 lL of diethyl ether on ice to the reaction tubes. Tubes were then transferred to 80 °C for 10 min to freeze the lower aqueous phase. Thereafter, the ether phase containing the steroids was discarded. The solution was then mixed thoroughly with 2:1 (vol/vol) of 0.25% dextran-coated (2.5%) charcoal and centrifuged at 2500g for 30 min at 4 °C to remove residual aromatics. A 500-lL volume of the supernatant was added in a scintillation vial containing 5 mL of scintillation cocktail, and 3H was measured as disintegrations per minute using a liquid scintillation counter. Aromatase activity was then expressed as femtomoles enzyme activity per hour per milligram of total protein content (fmol/h/mg protein). 2.6. Steroid hormone analysis Plasma concentrations of testosterone (T) and 17b-estradiol (E2) were measured using enzyme immunoassays (EIAs). Plasma was thawed on ice and steroids were extracted using organic solvent to prevent the interference of lipids and proteins in the analysis. Briefly, plasma was mixed with diethyl ether (4 the sample volume) by vortexing. After phase separation, the ether phase was transferred to a clean tube. This step was repeated three times.

Table 1 Primer pair sequences, accession numbers, amplicon size and annealing temperature conditions used for quantiative PCR. Target gene

Cyp19a Cyp19b Vitellogenin

50 -30 Primer sequence Forward

Reverse

GGGCACTGTCTGATGATGTC CTGACCCCTCTGGACACG AAGCCACCTCCAATGTCATC

GGGCTTGAGGAAGAACTCTG TCTCGTTGAGAGGCACCC GGGAGTCTGTCCCAAGACAA

Amplicon size (nucleotides)

Annealing temperature (oC)

GenBank Accession Nos.

104 96 391

60 55 60

DQ361037 DQ361038 DY802177

94

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

Finally, the ether was evaporated by heating to 30 °C. The dry extract was reconstituted in EIA buffer (Cayman) by vortexing. Enzyme immunoassay was performed according to the manufacturer’s protocol. Steroid hormone levels were quantified against a standard curve that was linearized using a logit transformation of B/B0 (bound sample/maximum bound). The sensitivity of the EIAs were T; 6 pg/ml and E2; 19 pg/ml. The intra- and inter-assay variations for both EIAs employed were 10% and 8%, respectively. Cross-reactivities between T and E2 EIAs and their respective steroid were 100%, and other steroids were all <0.01%. 2.7. Statistical analysis Statistical analyses were performed for all experiments (except for the immunochemical and immunohistochemical assays) using GraphPad Prism, version 5.00 (GraphPad Software Inc. 2007). Significant differences between control and exposure groups were performed using One-way ANOVA after testing for normality and variance homogeneity. Statistical differences between exposure groups were analyzed using the Tukey’s Multiple Comparison Test. The level of statistical significance was set at p < 0.05.

3. Results 3.1. Cellular distribution of neural aromatase immunoreactivity A saggital section of the Atlantic salmon parr brain stained with haemotoxylin-eosin is presented in Fig. 1. Immunohistochemical analyses demonstrated that aromatase protein was distributed throughout the salmon brain, with the strongest protein staining detected in the forebrain, including the telencephalon (Fig. 2A). Strong staining was also observed in all layers in the mesencephalon (Fig. 2B–C). In the cerebellum, immunoreactive aromatase protein was observed both in the molecular cell layer, cytoplasm of purkinje cells and throughout the granule cell layer (Fig. 3). Other brain regions exhibited lower immunoreactivity (micrographs not shown). No difference in staining intensity between individual fish and/or exposure groups could be observed. Control immunoreactive staining was performed by incubation with pre-immune serum showing elimination of protein labeling throughout the brain (Fig. 4A). Serial dilutions of the primary antibody produced progressive decreases in aromatase staining throughout the brain regions. Using Western blotting, immunodetectable aromatase protein was observed as a major band of 55 kDa, and minor differences (EE2 > NP5 > NP50), in protein band intensity was observed between control and exposed groups (Fig. 4B). 3.2. Cyp19 transcripts and aromatase activity levels

Fig. 1. Sagittal section of the Atlantic salmon brain (hematoxylin-eosin staining). Abbreviations: c, cerebellum; m, mesencephalon; mo, medulla oblongata; p, pituitary (hypophysis); sc, spinal cord; t, telencephalon. Scale bar: 2000 lm.

In the brain, mRNA level of Cyp19a was not affected by EE2 exposure, whereas 5 and 50 lg NP/L produced significant reductions of Cyp19a expression (Fig. 5A). For the classical brain isoform, Cyp19b, EE2 exposure produced significant increase of mRNA expression, showing a threefold elevation, compared to the solvent control (Fig. 5B). Fish exposed to NP concentrations did not produce any differences in brain Cyp19b expression, and remained at solvent control level (Fig. 5B). Brain aromatase activity was significantly reduced after exposure to NP concentrations (5 and 50 lg/L; Fig. 5C). No EE2 exposure related effect on brain aromatase activity was observed (Fig. 5C) and paralleled Cyp19a mRNA

Fig. 2. Aromatase immunoreactivity in the telencephalon (A) and mesencephalon (B and C [enlargement of framed field in B]) of immature Atlantic salmon. Micrographs are representative histological sections, as there were no differences in immunoreactivity between exposure groups. Abbreviations: m, mesencephalon; sgc, stratum griseum centrale; sp, stratum periventriculare; t, telencephalon. Scale bars: A, 250 lm; B, 500 lm; C, 100 lm.

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

95

Fig. 3. Aromatase immunoreactivity (A, C and D) and hematoxylin-eosin staining (B) of the cerebellum of immature Atlantic salmon. Image B–D are enlargements of framed fields 1–3 in image A. Micrographs are representative histological sections, as there were no differences in immunoreactivity between exposure groups. Abbreviations: c, capillaries; gl, granular layer; m, meninges; ml, molecular layer; p, purkinje cells. Scale bars: A, 500 lm; B, 220 lm; C, 150 lm; D, 100 lm.

pattern (Fig. 5A). For comparative purposes, absolute quantification of aromatase activity in unexposed fish showed the highest activity in brain (14.52 ± 1.86 fmol/h/mg protein, Fig. 5C) and liver (11.89 ± 1.55 fmol/h/mg protein; data not shown), and lower in head kidney (7.51 ± 1.37 fmol/h/mg protein, data not shown). 3.3. Plasma steroid, Vitellogenin mRNA and protein levels Fish exposed to EE2, 5 and 50 lg NP/L produced respective 980-, 1.5- and 620-fold increase of Vtg mRNA in the liver, compared to control (Fig. 6A). Additionally, we observed respective 6-, 1.2and 2.3-fold increase in plasma Vtg protein levels for fish exposed to EE2, 5 and 50 lg NP/L (Fig. 6B). Plasma T levels were 1.5 ± 0.5 ng/mL, and no differences were observed between the different exposure groups and the solvent control (Fig. 6C). In contrast, plasma levels of E2 significantly increased 2.9-, 2.3- and 3.9-fold when fish were treated with EE2, 5 and 50 lg NP/L, respectively, compared to control (Fig. 6D).

standing on the molecular mode of action of these chemicals and represent integral biological water quality criteria. Aromatase is considered to be an important potential EDC target, because modulation of aromatase expression, activity and function can dramatically alter the rate of estrogen production, impair fluctuating androgen/estrogen levels and thus may result in disruption of sex steroid hormone-related biological processes. Herein, we show the cellular localization and differences in expression patterns in the brain aromatase after exposure to EE2 and NP in relation to mRNA, protein and activity levels. We established that the previously observed xenoestrogen-induced alterations in brain Cyp19 and other steroidogenic genes paralleled modulation of functional protein levels, activity or designated to particular cellular localization in the brain. In addition, we also showed that these modulations paralleled alterations in plasma E2 levels, but not T. Overall, our data provide sensitive tools for evaluating water quality with added biological significance. 4.1. Distribution of brain aromatase protein immunoreactivity

4. Discussion Because of its widespread use, APEs represent an important class of ubiquitous chemicals that has been detected in the aquatic environment worldwide (Ahel et al., 1994). Several studies have identified NP as the most critical metabolite of APEs because of its enhanced resistance toward biodegradation, toxicity, estrogenic effects, and ability to bioaccumulate in aquatic organisms (Ahel et al., 1994). For example, NP has been shown to modulate brain aromatase activity and mRNA expression levels, in addition to other estrogenic effects in both in vivo and in vitro assays using a variety of teleost species (Arukwe et al., 1997b; Cheshenko et al., 2008; Jobling et al., 1996; Lyssimachou et al., 2006; Meucci and Arukwe, 2006b). Given that limited information exists concerning the potential disruption of the neuroendocrine systems by EDCs in general, and xenoestrogens in particular, the detection of Cyp19 mRNA expression and aromatase activity levels in fish after exposure to ubiquitous environmental contaminants will increase our under-

To our knowledge, this is the first study describing immunolocalization of aromatase protein in cell types in the Atlantic salmon brain. The distribution patterns were generally similar to what has been reported previously for other teleost species (Forlano et al., 2001; Gelinas and Callard, 1997; Marsh et al., 2006; Menuet et al., 2003; Pellegrini et al., 2005). Interestingly, we detected relatively strong aromatase protein immunostaining in the mesencephalon and cerebellum, which is in contrast to the relatively weak staining in these brain regions of another salmonid, the rainbow trout (Oncorhynchus mykiss) (Menuet et al., 2003). However, fully grown mature female rainbow trout were examined in the latter study, whereas the salmon examined in the present study were immature. This might explain the difference in staining patterns, as it is well established that mRNA and protein levels of sex steroid enzymes may show considerable variation due to sex, season, reproductive- and developmental stage of the individual fish. Although quantitative attempt was made to detect staining differences in the control and exposed samples, visual evaluation

96

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

Fig. 4. Control section (mesencephalon) incubated with pre-immune serum (A). Western blot (B) analysis of brain aromatase from immature Atlanitc salmon exposed to 17a-ethynylestradiol (EE2; 5 ng/L) and 4-nonylphenol (NP; 5 and 50 lg/ L). For each group, 10 lg protein per lane was applied in duplicates.

of the tissues did not produce any difference in staining intensity between control and exposure groups, and we acknowledge that this is partially in compliance with protein levels observed with immunoblotting. Nevertheless, it should be noted that the immunohistochemical techniques applied in this study was mainly used for qualitative purposes to detect tissue distribution after exposure to EDCs. In addition, the polyclonal antibody used in this study will not differentiate between aromatase isoforms. Thus, immunochemical assays will detect the total aromatase protein expression in a given tissue or cell. 4.2. Modulation of Cyp19 gene expression and enzyme activity Several studies have shown that Cyp19b (the dominant brain isoform) mRNA levels are up-regulated after E2 exposure (Gelinas et al., 1998; Lee et al., 2000; Tsai et al., 2000). Accordingly, we showed previously that estrogen-mimicking compounds, such as EE2 and NP generally increased brain Cyp19b mRNA levels (Lyssimachou et al., 2006; Meucci and Arukwe, 2006a). In the present

study, we expected a similar increasing pattern of Cyp19b transcription after EE2 and NP exposure. Indeed, we observed a threefold induction of Cyp19b transcription after 72 h exposure to 5 ng EE2/L, compared to control. Thus, the effect of EE2 on Cyp19b mRNA levels in the present study was even more evident than observed previously (Lyssimachou et al., 2006), where only the higher EE2 concentration (50 ng/L) produced a significant increase in Cyp19b mRNA. In contrast, we observed no effect of NP exposure (5 and 50 lg/L) on Cyp19b transcription levels. This is in compliance with our previous report (Meucci and Arukwe, 2006a), demonstrating only minor modulations of Cyp19b mRNA levels after 72 h NP exposure. However, the same study reported an evident concentration-dependent increase in Cyp19b mRNA levels after 7 days exposure to NP (5, 15 and 50 lg/L). These reports demonstrate that genes may show time-dependent expression patterns that need to be taken into consideration when predicting the effects of particular EDCs. Nevertheless, the general increasing pattern of Cyp19b transcription after exposure to estrogens and their mimics seems evident. Taken together, these and previous findings suggest that Cyp19b could be used as a sensitive exposure gene biomarker for environmentally relevant estrogenic compounds. On the contrary, mRNA expression of Cyp19a was not modulated after EE2 exposure, and Cyp19a mRNA levels were significantly reduced after NP exposure. In addition, brain Cyp19b mRNA levels were generally 10 times higher than Cyp19a levels. The differential abundance and expression of Cyp19 genes in the brain after exposure to estrogenic compounds have been reported previously (Kazeto et al., 2004; Lyssimachou et al., 2006; Meucci and Arukwe, 2006a; Trant et al., 2001). It has been speculated that these differences are a result of the Cyp19a lacking a definite physiological function in the brain in addition to an E2 negative feedback control mechanism (Meucci and Arukwe, 2006a). These speculations are supported by the fact that the teleost brain Cyp19a isoform lacks the consensus ERE in the promoter region, whereas the identification of EREs in the promoter region of the Cyp19b gene points to a direct estrogen responsiveness (Kazeto et al., 2001; Tchoudakova et al., 2001). Accordingly, the present and previous studies conducted in our laboratory suggest that environmental estrogens, including EE2 and NP, affect the Cyp19b in a similar way as natural estrogen.

Fig. 5. Transcriptional changes of brain Cyp19a (A), Cyp19b (B) and brain aromatase activity (C) of immature Atlantic salmon exposed to 17a-ethynylestradiol (EE2; 5 ng/L) and 4-nonylphenol (NP; 5 and 50 lg/L) for 72 h. Data are given as mean values expressed as percentage (%) of solvent control (n = 5) ± standard error of the mean (SEM). Asterisks denote exposure groups that are significantly different from the solvent control group (p < 0.05).

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

97

Fig. 6. Transcriptional changes of hepatic Vitellogenin (Vtg; A) and plasma levels Vitellogenin (B), testosterone (T; C) and estradiol-17b (E2; D) in immature Atlantic salmon exposed to 17a-ethynylestradiol (EE2; 5 ng/L) and 4-nonylphenol (NP; 5 and 50 lg/L) for 72 h. Data are given as mean values (n = 5) ± standard error of the mean (SEM). Asterisks denote exposure groups that are significantly different from the solvent control group (p < 0.05).

The aromatase activity was not significantly modulated after EE2 exposure, whereas both NP concentrations produced a decrease in aromatase activity. Comparing these results with a previous report from our group (Lyssimachou et al., 2006), it seems likely that the EE2 concentration employed in the present study (5 ng/L) was to low for increasing brain aromatase activity after 72 h exposure. Relatively little is known about direct xenoestrogenic effects on aromatase activity, but a few in vitro studies using mammalian cell lines have reported inhibition of aromatase activity by NP (Benachour et al., 2007; Bonefeld-Jorgensen et al., 2007). Additionally, reduced ovarian aromatase activity in the red mullet was suggested to be caused by NP (Martin-Skilton et al., 2006a). Interestingly, the decreased aromatase activity after NP exposure observed in the present study correlates perfectly with the transcription profile of brain Cyp19a. Indeed, the ovarian aromatase isoform (Cyp19a) has been shown to be more sensitive than the brain isoform to environmental estrogens such as flavones, but activities of the two aromatase isoforms seem to respond in a qualitatively similar manner (Zhao et al., 2001). Additionally, whereas brain aromatase activity was reduced after NP exposure, plasma levels of E2, the main aromatase metabolite, were increased. This indicates that local (brain) and systemic steroidogenic responses may show considerable variation after EDC exposure. 4.3. Modulation of sex steroid hormone plasma levels The production and secretion of steroid hormones are regulated by a complicated negative-feedback pathway that is turned on and off in response to fluctuating hormone levels. Accordingly, alterations in plasma sex steroid hormone concentrations may result from several different mechanisms of action, including direct effects on steroidogenic enzymes such as aromatase, or indirect modifications associated with altered feedback loops (Arukwe, 2001). It is well known that EDCs such as EE2 and NP can affect sex steroid hormone plasma levels in fish, but previous reports show inconsistent results. For example, juvenile Atlantic salmon injected with 1 and 5 mg/kg NP showed reduced E2 plasma levels, but a similar effect was not observed with higher (25 and 125 mg/ kg) doses (Arukwe et al. 1997a). Female rainbow trout (Oncorhynchus mykiss) exposed to waterborne NP (86 lg/L) showed decreased E2 plasma levels (Harris et al., 2001), both E2 and T

levels were reduced in juvenile turbot (Scophtalmus maximus) after waterborne NP exposure (30 lg/L), while juvenile Atlantic cod (Gadus morhua) were unaffected by the same exposure (Martin-Skilton et al., 2006b). However, Atlantic cod exposed to alkylphenols showed decreased E2 plasma levels in females (Meier et al., 2007), whereas octylphenol treatment (2, 20 and 200 mg/kg injections) resulted in increased plasma E2 levels in juvenile male summer flounder (Paralichthys dentatus) (Mills et al., 2001). Interestingly, although the fish in the present study were immature and had not developed distinct gonad tissue, we observed increased E2 levels after EE2 and NP treatment, whereas T levels were unaffected by the exposure. Thus, it is apparent that the steroiodogenic response may show considerable variation after exposure to sex steroid hormones and their mimics, depending on concentration, duration of the exposure, species, sex, cellular context and developmental stage of the individual fish. One possible explanation for this diversity in responses could be that sex steroid hormones and hormone mimics may affect other protein and enzymes involved in steroid metabolism. For example, we have previously observed that natural occurring steroid hormones and xenoestrogens (EE2 and NP) may produce alterations in transcription levels of the steroidogenic acute regulatory (StAR) protein and cytocrome P450-mediated side chain cleavage enzyme (P450scc) (Arukwe, 2005, 2008; Kortner and Arukwe, 2007; Kortner et al., 2009; Lyssimachou and Arukwe, 2007; Vang et al., 2007). The StAR protein and P450scc may be important molecular targets for different EDCs, as they mediate the rate-limiting step in acute steroid production (Stocco, 2001). However, StAR and P450scc generally show low expression levels, indicating a delicate regulation of steroidogenesis. 5. Conclusions The effects of EE2 and NP on the rate-limiting step in estrogen production were analyzed in immature Atlantic salmon brain, showing differential isoform-specific responses. Particularly, this study demonstrates that endocrine-disrupting effects of these two ubiquitous anthropogenic pollutants (EE2 and NP) may be observed on mRNA, protein and activity levels for the aromatase. Given that these effects partially paralleled cellular steroid hormone levels, it suggests a potential for adverse health and reproductive

98

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99

effects that may be used in the development of targeted test strategies for environmental risk assessment of pharmaceuticals and ubiquitous chemicals such as APEs, based on their mode(s) of action. Acknowledgments The authors thank Thomas Aurdal and Tora Bardal for technical assistance during sampling and analysis. This study was supported by the Norwegian research council. References Ahel, M., Giger, W., Koch, M., 1994. Behavior of alkylphenol polyethoxylate surfactants in the aquatic environment. 1. Occurrence and transformation in sewage-treatment. Water Research 28 (5), 1131–1142. Aherne, G.W., Briggs, R., 1989. The relevance of the presence of certain synthetic steroids in the aquatic environment. J. Pharm. Pharmacol. 41 (10), 735–736. Ahlborg, U.G., Lipworth, L., Titusernstoff, L., Hsieh, C.C., Hanberg, A., Baron, J., Trichopoulos, D., Adami, H.O., 1995. Organochlorine compounds in relation to breast-cancer, endometrial cancer, and endometriosis – an assessment of the biological and epidemiologic evidence. Crit. Rev. Toxicol. 25 (6), 463–531. Arukwe, A., 2001. Cellular and molecular responses to endocrine-modulators and the impact on fish reproduction. Mar. Pollut. Bull. 42 (8), 643–655. Arukwe, A., 2005. Modulation of brain steroidogenesis by affecting transcriptional changes of steroidogenic acute regulatory (StAR) protein and cholesterol side chain cleavage (P450scc) in juvenile Atlantic salmon (Salmo salar) is a novel aspect of nonylphenol toxicity. Environ. Sci. Technol, 39 (24), 9791–9798. Arukwe, A., 2008. Steroidogenic acute regulatory (StAR) protein and cholesterol side-chain cleavage (P450scc)-regulated steroidogenesis as an organ-specific molecular and cellular target for endocrine disrupting chemicals in fish. Cell Biol. Toxicol. 24 (6), 527–540. Arukwe, A., Förlin, L., Goksoyr, A., 1997a. Xenobiotic and steroid biotransformation enzymes in Atlantic Salmon (Samlo Salar) liver treated with an estrogenic compound, 4-nonylphenol. Environ. Tox. Chem. 16, 2576–2583. Arukwe, A., Knudsen, F.R., Goksoyr, A., 1997b. Fish zona radiata (eggshell) protein: a sensitive biomarker for environmental estrogens. Environ. Health Perspect. 105 (4), 418–422. Arukwe, A., Grotmol, T., Haugen, T.B., Knudsen, F.R., Goksoyr, A., 1999. Fish model for assessing the in vivo estrogenic potency of the mycotoxin zearalenone and its metabolites. Sci. Total Environ. 236 (1–3), 153–161. Arukwe, A., Thibaut, R., Ingebrigtsen, K., Celius, T., Goksoyr, A., Cravedi, J., 2000. In vivo and in vitro metabolism and organ distribution of nonylphenol in Atlantic salmon (Salmo salar). Aquat. Toxicol. 49 (4), 289–304. Baulieu, E.E., Robel, P., Schumacher, M., 2001. Neurosteroids: beginning of the story. Int. Rev. Neurobiol. 46, 1–32. Benachour, N., Moslemi, S., Sipahutar, H., Seralini, G.E., 2007. Cytotoxic effects and aromatase inhibition by xenobiotic endocrine disrupters alone and in combination. Toxicol. Appl. Pharmacol. 222 (2), 129–140. Blackburn, M.A., Waldock, M.J., 1995. Concentrations of alkylphenols in rivers and Estuaries in England and Wales. Water Research 29 (7), 1623–1629. Bonefeld-Jorgensen, E.C., Long, M., Hofmeister, M.V., Vinggaard, A.M., 2007. Endocrine-disrupting potential of bisphenol A, bisphenol A dimethacrylate, 4n-nonylphenol, and 4-n-octylphenol in vitro: new data and a brief review. Environ. Health Perspect. 115 (Suppl. 1), 69–76. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein–dye binding. Anal. Biochem. 72, 248–254. Caldwell, J., 1986. Conjugation mechanisms of xenobiotic metabolism – mammalian aspects. Acs Symposium Series 299, 2–28. Callard, G.V., Tchoudakova, A.V., Kishida, M., Wood, E., 2001. Differential tissue distribution, developmental programming, estrogen regulation and promoter characteristics of cyp19 genes in teleost fish. J. Steroid Biochem. Mol. Biol. 79 (1–5), 305–314. Celius, T., Haugen, T.B., Grotmol, T., Walther, B.T., 1999. A sensitive zonagenetic assay for rapid in vitro assessment of estrogenic potency of xenobiotics and mycotoxins. Environ. Health Perspect. 107 (1), 63–68. Cheshenko, K., Pakdel, F., Segner, H., Kah, O., Eggen, R.I., 2008. Interference of endocrine disrupting chemicals with aromatase CYP19 expression or activity, and consequences for reproduction of teleost fish. Gen. Comp. Endocrinol. 155 (1), 31–62. Ekelund, R., Bergman, A., Granmo, A., Berggren, M., 1990. Bioaccumulation of 4nonylphenol in marine animals – a reevaluation. Environ. Poll. 64 (2), 107–120. Forlano, P.M., Deitcher, D.L., Myers, D.A., Bass, A.H., 2001. Anatomical distribution and cellular basis for high levels of aromatase activity in the brain of teleost fish: aromatase enzyme and mRNA expression identify glia as source. J. Neurosci. 21 (22), 8943–8955. Forlano, P.M., Schlinger, B.A., Bass, A.H., 2006. Brain aromatase: new lessons from non-mammalian model systems. Front. Neuroendocrinol. 27 (3), 247–274. Gelinas, D., Callard, G.V., 1997. Immunolocalization of aromatase- and androgen receptor-positive neurons in the goldfish brain. Gen. Comp. Endocrinol. 106 (2), 155–168.

Gelinas, D., Pitoc, G.A., Callard, G.V., 1998. Isolation of a goldfish brain cytochrome P450 aromatase cDNA: mRNA expression during the seasonal cycle and after steroid treatment. Mol. Cell. Endocrinol. 138 (1–2), 81–93. Guardans, R., Gimeno, B.S., 1994. Long-distance transportation of atmospheric pollutants and its effects on ecosystems. Microbiologia 10 (1–2), 145–158. Harris, C.A., Santos, E.M., Janbakhsh, A., Pottinger, T.G., Tyler, C.R., Sumpter, J.P., 2001. Nonylphenol affects gonadotropin levels in the pituitary gland and plasma of female rainbow trout. Environ. Sci. Technol. 35 (14), 2909– 2916. Jobling, S., Sheahan, D., Osborne, J.A., Matthiessen, P., Sumpter, J.P., 1996. Inhibition of testicular growth in rainbow trout (Oncorhynchus mykiss) exposed to estrogenic alkylphenolic chemicals. Environ. Tox. Chem. 15 (2), 194–202. Kazeto, Y., Ijiri, S., Place, A.R., Zohar, Y., Trant, J.M., 2001. The 50 -flanking regions of CYP19A1 and CYP19A2 in zebrafish. Biochem. Biophys. Res. Commun. 288 (3), 503–508. Kazeto, Y., Place, A.R., Trant, J.M., 2004. Effects of endocrine disrupting chemicals on the expression of CYP19 genes in zebrafish (Danio rerio) juveniles. Aquat. Toxicol. 69 (1), 25–34. Kishida, M., McLellan, M., Miranda, J.A., Callard, G.V., 2001. Estrogen and xenoestrogens upregulate the brain aromatase isoform (P450aromB) and perturb markers of early development in zebrafish (Danio rerio). Comp. Biochem. Physiol. B Biochem. Mol. Biol. 129 (2–3), 261–268. Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber, L.B., Buxton, H.T., 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in US streams, 1999–2000: a national reconnaissance. Environ. Sci. Technol. 36 (6), 1202–1211. Kortner, T.M., Arukwe, A., 2007. The xenoestrogen, 4-nonylphenol, impaired steroidogenesis in previtellogenic oocyte culture of Atlantic cod (Gadus morhua) by targeting the StAR protein and P450scc expressions. Gen. Comp. Endocrinol. 150 (3), 419–429. Kortner, T.M., Rocha, E., Arukwe, A., 2009. Previtellogenic oocyte growth and transcriptional changes of steroidogenic enzyme genes in immature female Atlantic cod (Gadus morhua L.) after exposure to the androgens 11ketotestosterone and testosterone. Comp. Biochem. Physiol. A Mol. Integr. Physiol. 152 (3), 304–313. Kvestak, R., Terzic, S., Ahel, M., 1994. Input and distribution of alkylphenol polyethoxylates in a stratified estuary. Mar. Chem. 46 (1–2), 89–100. Larsson, D.G.J., Adolfsson-Erici, M., Parkkonen, J., Pettersson, M., Berg, A.H., Olsson, P.E., Forlin, L., 1999. Ethinyloestradiol – an undesired fish contraceptive? Aquat. Toxicol. 45 (2–3), 91–97. Lee, Y.H., Lee, F.Y., Yueh, W.S., Tacon, P., Du, J.L., Chang, C.N., Jeng, S.R., Tanaka, H., Chang, C.F., 2000. Profiles of gonadal development, sex steroids, aromatase activity, and gonadotropin II in the controlled sex change of protandrous black porgy, Acanthopagrus schlegeli Bleeker. Gen. Comp. Endocrinol. 119 (1), 111– 120. Lyssimachou, A., Arukwe, A., 2007. Alteration of brain and interrenal StAR protein, P450scc, and Cyp11beta mRNA levels in Atlantic salmon after nominal waterborne exposure to the synthetic pharmaceutical estrogen ethynylestradiol. J. Toxicol. Environ. Health A 70 (7), 606–613. Lyssimachou, A., Jenssen, B.M., Arukwe, A., 2006. Brain cytochrome P450 aromatase gene isoforms and activity levels in Atlantic salmon after waterborne exposure to nominal environmental concentrations of the pharmaceutical ethynylestradiol and antifoulant tributyltin. Toxicol. Sci. 91 (1), 82–92. Marsh, K.E., Creutz, L.M., Hawkins, M.B., Godwin, J., 2006. Aromatase immunoreactivity in the bluehead wrasse brain, Thalassoma bifasciatum: immunolocalization and co-regionalization with arginine vasotocin and tyrosine hydroxylase. Brain Res. 1126 (1), 91–101. Martin-Skilton, R., Lavado, R., Thibaut, R., Minier, C., Porte, C., 2006a. Evidence of endocrine alteration in the red mullet, Mullus barbatus from the NW Mediterranean. Environ. Pollut. 141 (1), 60–68. Martin-Skilton, R., Thibaut, R., Porte, C., 2006b. Endocrine alteration in juvenile cod and turbot exposed to dispersed crude oil and alkylphenols. Aquat. Toxicol. 78 (Suppl. 1), S57–S64. Meier, S., Andersen, T.E., Norberg, B., Thorsen, A., Taranger, G.L., Kjesbu, O.S., Dale, R., Morton, H.C., Klungsoyr, J., Svardal, A., 2007. Effects of alkylphenols on the reproductive system of Atlantic cod (Gadus morhua). Aquat. Toxicol. 81 (2), 207–218. Menuet, A., Anglade, I., Le Guevel, R., Pellegrini, E., Pakdel, F., Kah, O., 2003. Distribution of aromatase mRNA and protein in the brain and pituitary of female rainbow trout: comparison with estrogen receptor alpha. J. Comp. Neurol. 462 (2), 180–193. Menuet, A., Pellegrini, E., Brion, F., Gueguen, M.M., Anglade, I., Pakdel, F., Kah, O., 2005. Expression and estrogen-dependent regulation of the zebrafish brain aromatase gene. J. Comp. Neurol. 485 (4), 304–320. Meucci, V., Arukwe, A., 2006a. Transcriptional modulation of brain and hepatic estrogen receptor and P450arom isotypes in juvenile Atlantic salmon (Salmo salar) after waterborne exposure to the xenoestrogen, 4-nonylphenol. Aquat. Toxicol. 77 (2), 167–177. Meucci, V., Arukwe, A., 2006b. The xenoestrogen 4-nonylphenol modulates hepatic gene expression of pregnane X receptor, aryl hydrocarbon receptor, CYP3A and CYP1A1 in juvenile Atlantic salmon (Salmo salar). Comp. Biochem. Physiol. C Toxicol. Pharmacol. 142 (1–2), 142–150. Mills, L.J., Gutjahr-Gobell, R.E., Haebler, R.A., Horowitz, D.J., Jayaraman, S., Pruell, R.J., McKinney, R.A., Gardner, G.R., Zaroogian, G.E., 2001. Effects of estrogenic (o,p’DDT; octylphenol) and anti-androgenic (p,p’-DDE) chemicals on indicators of

T.M. Kortner et al. / General and Comparative Endocrinology 164 (2009) 91–99 endocrine status in juvenile male summer flounder (Paralichthys dentatus). Aquat. Toxicol. 52 (2), 157–176. Mortensen, A.S., Arukwe, A., 2007. Effects of 17alpha-ethynylestradiol on hormonal responses and xenobiotic biotransformation system of Atlantic salmon (Salmo salar). Aquat. Toxicol. 85 (2), 113–123. Nash, J.P., Kime, D.E., Van der Ven, L.T., Wester, P.W., Brion, F., Maack, G., Stahlschmidt-Allner, P., Tyler, C.R., 2004. Long-term exposure to environmental concentrations of the pharmaceutical ethynylestradiol causes reproductive failure in fish. Environ. Health Perspect. 112 (17), 1725–1733. Naylor, C.G., Mieure, J.P., Adams, W.J., Weeks, J.A., Castaldi, F.J., Ogle, L.D., Romano, R.R., 1992. Alkylphenol ethoxylates in the environment. J. Am. Oil Chem. Soc. 69 (7), 695–703. Pasmanik, M., Callard, G.V., 1985. Aromatase and 5 alpha-reductase in the teleost brain, spinal cord, and pituitary gland. Gen. Comp. Endocrinol. 60 (2), 244–251. Pelissero, C., Bennetau, B., Babin, P., Le Menn, F., Dunogues, J., 1991. The estrogenic activity of certain phytoestrogens in the Siberian sturgeon Acipenser baeri. J. Steroid Biochem. Mol. Biol. 38 (3), 293–299. Pelissero, C., Flouriot, G., Foucher, J.L., Bennetau, B., Dunogues, J., Le Gac, F., Sumpter, J.P., 1993. Vitellogenin synthesis in cultured hepatocytes; an in vitro test for the estrogenic potency of chemicals. J. Steroid Biochem. Mol. Biol. 44 (3), 263–272. Pellegrini, E., Menuet, A., Lethimonier, C., Adrio, F., Gueguen, M.M., Tascon, C., Anglade, I., Pakdel, F., Kah, O., 2005. Relationships between aromatase and estrogen receptors in the brain of teleost fish. Gen. Comp. Endocrinol. 142 (1–2), 60–66. Rotchell, J.M., Ostrander, G.K., 2003. Molecular markers of endocrine disruption in aquatic organisms. J. Toxicol. Environ. Health B Crit. Rev. 6 (5), 453–496.

99

Schumacher, M., Robel, P., Baulieu, E.E., 1996. Development and regeneration of the nervous system: a role for neurosteroids. Dev. Neurosci. 18 (1–2), 6–21. Segner, H., Caroll, K., Fenske, M., Janssen, C.R., Maack, G., Pascoe, D., Schafers, C., Vandenbergh, G.F., Watts, M., Wenzel, A., 2003. Identification of endocrinedisrupting effects in aquatic vertebrates and invertebrates: report from the European IDEA project. Ecotoxicol. Environ. Saf. 54 (3), 302–314. Simpson, E.R., Clyne, C., Rubin, G., Boon, W.C., Robertson, K., Britt, K., Speed, C., Jones, M., 2002. Aromatase – a brief overview. Annu. Rev. Physiol. 64, 93–127. Stocco, D.M., 2001. StAR protein and the regulation of steroid hormone biosynthesis. Annu. Rev. Physiol. 63, 193–213. Tchoudakova, A., Kishida, M., Wood, E., Callard, G.V., 2001. Promoter characteristics of two cyp19 genes differentially expressed in the brain and ovary of teleost fish. J. Steroid Biochem. Mol. Biol. 78 (5), 427–439. Trant, J.M., Gavasso, S., Ackers, J., Chung, B.C., Place, A.R., 2001. Developmental expression of cytochrome P450 aromatase genes (CYP19a and CYP19b) in zebrafish fry (Danio rerio). J. Exp. Zool. 290 (5), 475–483. Tsai, C.L., Wang, L.H., Chang, C.F., Kao, C.C., 2000. Effects of gonadal steroids on brain serotonergic and aromatase activity during the critical period of sexual differentiation in tilapia, Oreochromis mossambicus. J. Neuroendocrinol. 12 (9), 894–898. Vang, S.H., Kortner, T.M., Arukwe, A., 2007. Steroidogenic acute regulatory (StAR) protein and cholesterol side-chain cleavage (P450scc) as molecular and cellular targets for 17alpha-ethynylestradiol in salmon previtellogenic oocytes. Chem. Res. Toxicol. 20 (12), 1811–1819. Zhao, J., Mak, P., Tchoudakova, A., Callard, G., Chen, S., 2001. Different catalytic properties and inhibitor responses of the goldfish brain and ovary aromatase isozymes. Gen. Comp. Endocrinol. 123 (2), 180–191.