Effects of selected xenoestrogens on liver peroxisomes, vitellogenin levels and spermatogenic cell proliferation in male zebrafish

Comparative Biochemistry and Physiology, Part C 141 (2005) 133 – 144 www.elsevier.com/locate/cbpc

Effects of selected xenoestrogens on liver peroxisomes, vitellogenin levels and spermatogenic cell proliferation in male zebrafishB Maren Ortiz-Zarragoitia, Miren P. Cajaraville* Biologia Zelularra eta Histologia Laborategia, Zoologia eta Biologia Zelularra Saila, Zientzia eta Teknologia Fakultatea, UPV/EHU, 644 PK, E-48080 Bilbo, Basque Country, Spain Received 7 December 2004; received in revised form 30 April 2005; accepted 2 May 2005 Available online 5 July 2005

Abstract Environmental estrogenic compounds or xenoestrogens can mimic natural estrogens and cause a variety of adverse effects on aquatic wildlife. The purpose of the present work was to investigate if xenoestrogens are able to cause proliferation of liver peroxisomes using zebrafish (Danio rerio) as a model. Adult male zebrafish were exposed for 15 days to 17b-estradiol (E2) and the xenoestrogens dibutylphthalate (DBP), methoxychlor (MXC), 4-tert-octylphenol (OP) and 17a-ethynylestradiol (EE2). All five tested compounds caused significant proliferation of liver peroxisomes ( p < 0.05) as indicated by increased peroxisomal surface and numerical densities and elevated activities of the peroxisomal b-oxidation enzyme acyl-CoA oxidase (AOX). In the case of DBP, MXC and E2, positive significant correlations between peroxisomal density parameters and AOX were found. The treatments did not produce gross alterations in testis histology, but spermatogenic cell proliferation was disturbed in E2 and EE2-treated groups and vitellogenin levels increased significantly in fish exposed to MXC, OP, EE2 and E2 with respect to controls. Furthermore, a significant correlation between vitellogenin levels and AOX activity was found for MXC, OP and EE2 treatments, suggesting that for the latter xenoestrogens early estrogenic effects are associated with liver peroxisome proliferation. No such association occurred with typical peroxisome proliferators such as DBP. D 2005 Elsevier Inc. All rights reserved. Keywords: Dibutylphthalate; Methoxychlor; 4-tert-octylphenol; 17a-ethynylestradiol; 17b-estradiol; Spermatogenesis; Liver peroxisome proliferation; Vitellogenin; Zebrafish

1. Introduction Several chemicals present in the aquatic environment are able to interfere with the endocrine system causing adverse effects on growth, behavior, reproduction and immune function of aquatic wildlife. Those chemicals that alter function(s) of the endocrine system and consequently cause adverse health effects in an intact organism, or its progeny, or (sub)populations have been named ‘‘endocrine-disrupting chemicals’’ (WHO/IPCS, 2002). Xenoestrogens constitute a particular class of endocrine disruptors with the ability to i Presented in part at the 11th Annual Meeting of the Society of Environmental Toxicology and Chemistry-Europe, Madrid, Spain, May 6 – 10, 2001. * Corresponding author. Tel.: +34 94 6012697; fax: +34 94 6013500. E-mail address: [email protected] (M.P. Cajaraville).

1532-0456/$ - see front matter D 2005 Elsevier Inc. All rights reserved. doi:10.1016/j.cca.2005.05.010

mimic natural estrogens. These include certain pesticides, detergent derivatives, antioxidants, plasticizers, PCB congeners and many other industrial chemicals (WHO/IPCS, 2002; Goksøyr et al., 2003). Xenoestrogens may elicit effects through a number of divergent pathways including direct binding and activation of the estrogen receptor (ER); binding to other nuclear receptors which interact with an estrogen response element (ERE); and through other receptor and/or signal transduction pathways (Gillesby and Zacharewski, 1998; Arukwe and Goksøyr, 2003; Kirk et al., 2003). The ER belongs to a structurally related family of proteins called the nuclear receptor superfamily, which also includes the peroxisome proliferator-activated receptors (PPARs) (Barton and Andersen, 1998). PPARs, as ER, bind to a special response element on target DNA, named peroxisome proliferator response element (PPRE) (Lemberger et al.,

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1996). Interactions between these receptors and their response elements in DNA have been demonstrated. PPAR forms heterodimers with the retinoid-X-receptor (RXR) and the dimer has been shown to bind to ERE, thus activating genes normally controlled by estrogen (Lemberger et al., 1996; Gillesby and Zacharewski, 1998; Keller et al., 2000). It is also possible that estrogen receptors activate PPAR target genes via PPRE or ERE (Djouadi et al., 1998), maybe after estrogen-induced production of metabolites capable of activating PPARs (Ma et al., 1998). Modulation of PPARcinduced genes by ER has been demonstrated in E2 exposed cells (Wang and Kilgore, 2002). All these data point to a possible relationship between peroxisome proliferation responses and estrogenic effects. ER shows capacity to bind different estrogen analogues as well as structurally diverse non-steroidal compounds (Crews et al., 2000). In the same way, a range of xenobiotic compounds interact with PPAR inducing peroxisome proliferation (Cajaraville et al., 2003a). Many enzymes present in liver peroxisomes are associated with lipid metabolism, namely with b-oxidation of very long-chain fatty acids and biologically important lipid derivatives such as prostaglandins and leukotrienes. Peroxisomes play a role in the synthesis of plasmalogens and cholesterol, the precursor of steroid hormones. They are also involved in the degradation of steroids as the key enzyme for the inactivation of E2, 17b-hydroxyesteroid dehydrogenase, is localized in peroxisomes (Markus et al., 1995; Keller et al., 2000). Furthermore, peroxisomes are present in steroidogenic tissues in mammals where an important role of peroxisomes in steroid biosynthesis has been proposed (Magalhaˆes and Magalhaˆes, 1997). Peroxisome proliferation is a pleiotropic cellular response which involves a drastic increase in peroxisomal volume and induction of boxidation enzymes, caused by several factors including exposure to organic environmental pollutants (Cancio and Cajaraville, 2000; Cajaraville et al., 2003a). Some peroxisome proliferators such as phthalate ester plasticizers have been identified as weak estrogens in several in vitro tests (Jobling et al., 1995; Harris et al., 1997). The aim of the present work was to investigate if xenoestrogens are able to cause proliferation of liver peroxisomes in a series of short-term experiments with zebrafish (Danio rerio). The activity of the key peroxisomal b-oxidation enzyme, acyl-CoA oxidase (AOX), and peroxisomal density parameters were used as markers for peroxisome proliferation (Cajaraville et al., 2000, 2003a; Cancio and Cajaraville, 2000). Vitellogenin levels, extensively used to assess exposure to estrogenic compounds in aquatic animals (Matthiessen and Sumpter, 1998; Goksøyr et al., 2003), were measured as a marker for estrogenic effects. Additionally, testis histology and spermatogenic cell proliferation were assessed to study any possible effects of tested compounds on gonad morphology and gamete development. Zebrafish are highly sensitive to estrogens when exposure takes place in early life stages, and it has

been suggested that zebrafish are ideal model organisms for the evaluation of estrogenic effects (Arukwe and Goksøyr, ¨ rn et al., 2003; Segner et al., 2003). In addition, we 2003; O have recently demonstrated that the three PPAR subtypes, PPARa, PPARb and PPARc are expressed in different tissues of adult zebrafish (Ibabe et al., 2002). In the present work, adult male zebrafish were exposed to 17b-estradiol (E2) and to known or suspected estrogenic pollutants such as the phthalate dibutylphthalate (DBP), the organochlorine pesticide methoxychlor ((1,1,1-tri-chloro-2,2-bis(4-methoxyphenil)-ethane), MXC), the alkylphenol ethoxylate derivative 4-tert-octylphenol (OP) and the synthetic estrogen 17a-ethynylestradiol (EE2). DBP was dosed at a concentration of 500 Ag/L as this dose has been shown to induce peroxisome proliferation in previous studies in molluscs (Cancio et al., 1998). Reported maximum concentrations of DBP in developing countries range from 10 to 1472 mg/L (Fatoki and Ogunfowokan, 1993), and typical concentrations in the western world are 0.3 –30 Ag/L (Tyler et al., 1998). MXC was tested at a dose of 100 Ag/L, below 96 h LC50 values reported for several teleost species (10 to 260 Ag/L) (Krisfalusi et al., 1998). OP was tested at a concentration of 500 Ag/L, as concentrations up to 600 Ag/L have been detected in more polluted sewage effluents (Tyler et al., 1998; Sole´ et al., 2000a). Finally, for EE2 and E2 we selected a dose (10 Ag/L) close to that known to cause peroxisome proliferation and estrogenic effects in adult male zebrafish (Veranicˆ and Pipan, 1992). This dose was within the sublethal range based on reported 72 h LC50 values of E2 for Japanese medaka (Kashiwada et al., 2001).

2. Materials and methods 2.1. Animals and treatments Adult male zebrafish (Danio rerio), obtained from a petshop (ABERIAK, Getxo), were maintained in 15 L aquaria at a constant temperature of 26 -C, pH 7.6 and 7.46 mg/L dissolved oxygen. The light cycle was maintained at 14 h light and 10 h dark. Sixty fish in two replicates per experimental group were exposed for 15 days to DBP,1 MXC,2 OP,3 EE24 and E2.5 Dimethylformamide (DMF, 0.006 ml/L)6 was used as vehicle, and thus a DMF control group was run in parallel. The doses used during experiments were 500 Ag/L for DBP, 100 Ag/L for MXC, 500 Ag/ L for OP and 10 Ag/L for EE2 and E2. Water was renewed every day and chemicals were added after each change of water. Specimens were fed to satiation every day during the 1 2 3 4 5 6

DBP obtained from Fluka, Buchs, Switzerland, 98% GC. MXC obtained from Riedel-de Hae¨n, Seelze, Germany, 98% HPLC. OP obtained from Fluka, Buchs, Switzerland, 90% GC. EE2 obtained from Fluka, Buchs, Switzerland, 85% HPLC. E2 obtained from Fluka, Buchs, Switzerland, 97% HPLC. DMF obtained from Merck, Darmstadt, Germany,  99% GC.

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experiment with commercial JBL-Gala flakes. Fish were anaesthetized with a saturated solution of 3-aminobenzoic acid ethyl ester (Sigma, St Louis, MO, USA) and then sacrificed after 7 and 15 days of exposure. 2.2. Enzyme activity measurements Samples of whole visceral mass (digestive tract plus surrounding liver) were processed for spectrophotometric measurement of activities of the peroxisomal enzymes catalase and acyl-CoA oxidase (AOX). Visceral mass pieces from 18– 20 fish per exposure group were removed and placed in ice-cold 60 mM Tris, 0.25 M sucrose buffer (pH 8.3), homogenized in 4 subgroups and centrifuged at 600 g at 4 -C for 20 min. Peroxisomal AOX activity was measured in duplicate supernatants using a Shimadzu UV-1603 spectrophotometer (Duisburg, Germany) by monitoring the H2O2-dependent dehydrogenation of 0.05 mM 2V, 7V-dichlorodihydrofluorescein diacetate (Molecular Probes; Eugene, Oregon, USA) at 502 nm in the presence of horseradish peroxidase (12 Units/ml) and using 30 AM palmitoyl-CoA as substrate in 0.01 M potassium-phosphate buffer (Small et al., 1985). The activity is given as mU AOX mg 1 protein equivalent to nmol H2O2 min 1 mg 1 protein. Catalase activity was determined by measuring the decomposition of hydrogen peroxide at 240 nm (Aebi, 1974). Catalase activity is given as mmol H2O2 min 1 mg 1 protein. Protein concentration was determined by the DC Protein Assay of Bio-Rad (Hercules, California, USA). 2.3. Vitellogenin measurements Vitellogenin levels were determined using a commercial EIA kit for zebrafish vitellogenin (Biosense Laboratories SA, Bergen, Norway) in the same homogenates used for enzyme activity measurements (3 replicates per each experimental group), following the instructions of the manufacturer. Quantification was performed in two pools of samples at 490 nm and results are given as Ag vitellogenin per mg protein. 2.4. Liver peroxisome histochemistry and stereology Five livers from each experimental group were dissected out and processed for catalase histochemistry according to the protocol of Braunbeck et al. (1990). Briefly, livers were fixed in 2.5% glutaraldehyde in 0.1 M sodium cacodylate buffer (pH 7.6) containing 4% polyvinylpyrrolidone and 0.05% calcium chloride for 2  20 min at 4 -C. After rinsing in cacodylate buffer, samples were incubated in 10 mM Teorell– Stenhagen buffer (pH 10) containing 5 mM 3,3’diaminobenzidine and 0.5% H2O2 for 60 min at 37 -C in a shaking water bath. Subsequently, samples were rinsed with Teorell– Stenhagen buffer and post-fixed for 60 min at 4 -C with 1% osmium ferricyanide (1 : 1 mixture of 2% aqueous OsO4 and 2% K4(Fe(CN)6)). After rinsing in cacodylate

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buffer, samples were dehydrated in a graded series of ethanol and embedded in Epon 812. Semi-thin sections of 1 Am were cut using a Leica Ultracut UCT ultramicrotome (Vienna, Austria), mounted in DPX (Fluka, Buchs, Switzerland) and examined under a Leitz Laborlux S light microscope (Wetzlar, Germany). For the stereological analysis of peroxisomes, a lattice with 168 test points (Weibel’s multipurpose test system p168) was used following the method described in Cajaraville et al. (1997). Four parameters were calculated: peroxisomal volume density (V VP = V P/V C), peroxisomal surface density (S VP = S P/V C), peroxisomal surface to volume ratio (SV P = S P/V P) and peroxisomal numerical density (N VP = N P/V C), where V=volume, S=surface, N=number, P=peroxisome and C=hepatocyte cytoplasm. 2.5. Testis histology and immunohistochemistry of PCNA Samples of testis were dissected out, fixed in neutral buffered formalin (10%) and routinely processed for paraffin embedding. Sections (7 Am thick) were cut in a Leitz 1512 microtome (Vienna, Austria) and stained with hematoxylin/eosin for the observation of general histology. Proliferation of spermatogenic cells was analyzed by image analysis in 3 Am sections of paraffin-embedded samples processed for the immunohistochemical detection of the proliferating cell nuclear antigen (PCNA). Sections mounted in silanized slides were dried at 60 -C for 20 min. Then, sections were de-paraffinized in xylene, hydrated in a graded series of ethanol and brought to distilled water. All samples were subjected to antigen retrieval in Tris buffered saline (TBS) containing 0.1% trypsin for 5 min at 37 -C. After rinsing in TBS, endogenous peroxidase activity was blocked by incubating the sections in methanol containing 3% H2O2 for 10 min. Sections were then washed in TBS and incubated with avidin dissolved in TNB (NEN, Boston, USA) (1 : 8) for 60 min. After a brief rinse in TBS, sections were incubated overnight with a mouse monoclonal antibody against PCNA, clone PC10 (Sigma, St. Louis, MO, USA), diluted 1 : 200 in a solution of biotin in TNB (1 : 8) containing 0.05% Tween 20. Control sections were incubated without the antibody. After several rinses in TBS, immunocomplexes were visualized with the avidin-biotin-enzyme complex (ABC) method, using the Peroxidase Vectastain Elite ABC Kit (Burlingame, California, USA) as follows. Sections were incubated for 60 min with biotinylated horse anti-mouse IgG secondary antibody supplied in the kit. After several rinses in TBS, sections were incubated for 25 min with ABC reagent also supplied in the kit. After a brief rinse in TBS, the visualization of peroxidase activity was achieved using a chromogen solution containing 3-amino-9-ethyl-carbazole (AEC) (Sigma, St. Louis, MO, USA) (0.5 mg/ml), H2O2 (0.03%) and 0.05 M sodium acetate buffer, pH 5.2. Finally, after a brief rinse in distilled water, sections were mounted in Kaiser’s glycerol gelatin. Sections were examined under

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a Leitz Laborlux S light microscope (Wetzlar, Germany) and the surface density of PCNA positive nuclei was measured by image analysis using an objective lens of 100. Based on the different size of positive nuclei, two types of PCNA positive cells were identified (spermatogo-

nial and spermatocyte cells), and the number of each type of cell per unit of gonad area was measured using the Weibel’s multipurpose test system p168. In each case five fields per animal were measured in a total of five animals per experimental group.

Fig. 1. Surface density (S VP) and numerical density (N VP) of liver peroxisomes in adult male zebrafish treated for 7 and 15 days with dimethylformamide used as solvent (solid bars) and with dibutylphthalate (DBP), methoxychlor (MXC), 4-tert-octylphenol (OP), 17a-ethynylestradiol (EE2) and 17b-estradiol (E2) (open bars). Samples of livers of fish exposed to OP for 15 days were lost. Data are represented as mean T standard deviations. Asterisks indicate statistically significant differences ( p < 0.05) between solvent control and exposed groups within each exposure duration.

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2.6. Statistics Statistical analyses were performed with the aid of the SPSS/PC + statistical package (SPSS Inc., Redmond, WA). Significant differences between control and exposed groups were studied using the Student’s t-test, after testing for

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normality of the data and homogeneity of variances. When data did not follow these assumptions they were logtransformed before statistical analysis. Significance was established at p < 0.05. Possible correlation between different parameters was investigated using Spearman’s correlation index.

Fig. 2. Liver acyl-CoA oxidase (AOX) and catalase activities in adult male zebrafish treated for 7 and 15 days with dimethylformamide used as solvent (solid bars) and with dibutylphthalate (DBP), methoxychlor (MXC), 4-tert-octylphenol (OP), 17a-ethynylestradiol (EE2) and 17b-estradiol (E2) (open bars). Data are represented as mean T standard deviations. Asterisks indicate statistically significant differences ( p < 0.05) between solvent control and exposed groups within each exposure duration.

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3. Results After diaminobenzidine staining, peroxisomes were easily identified as brown-black spherical structures in liver hepatocytes. The volume, surface and numerical densities of peroxisomes changed significantly in exposed zebrafish when compared to controls (Fig. 1) while few differences in peroxisomal surface to volume ratios were found (not shown). As changes in peroxisomal volume and surface densities followed the same pattern of variation only surface densities are given in Fig. 1 together with numerical densities. DBP caused a significant increase in peroxisomal surface density after 7 and 15 days of exposure, together with a higher numerical density of peroxisomes after 15 days (Fig. 1). MXC exposure provoked significant increases in peroxisomal surface and numerical densities at both sampling periods (Fig. 1). In the case of OP significant increases in surface and numerical densities were detected at day 7 of exposure (Fig. 1). No data is available for the second (15 days) sampling because those samples were lost. Zebrafish exposed to EE2 showed a significant increase of both surface and numerical densities after 7 days exposure and the same result was obtained with E2 after 15 days exposure (Fig. 1). The reported increases in peroxisomal densities were accompanied by increases in the activity of the peroxisomal enzyme acyl-CoA oxidase (Fig. 2). This enzyme activity was significantly elevated in zebrafish exposed to DBP and MXC at both sampling times, in those exposed to OP only after 15 days of exposure, and in fish exposed to E2 and EE2 only after 7 days of exposure. The correlation analysis confirmed the association between peroxisomal density parameters and the activity of acyl-CoA oxidase (Table 1), significant positive correlation coefficients being found in experiments with DBP, MXC and E2. No clear trends were observed for variations in the activity of the other peroxisomal enzyme studied, catalase (Fig. 2). Catalase activity did not change in fish exposed to DBP, OP or E2 but increased significantly in zebrafish exposed to MXC for 15 days and was reduced significantly in those exposed to EE2 for 7 days. Table 1 Correlation indices between peroxisomal stereological parameters and acylCoA oxidase (AOX) activities in different experiments where adult male zebrafish were treated with dibutylphthalate (DBP), methoxychlor (MXC), 17a-ethynylestradiol (EE2) and 17b-estradiol (E2) Stereological parameters

DBP

MXC

EE2

E2

AOX activity Volume density Surface density Surface to volume ratio Numerical density

0.7008 0.9347* 0.4455 0.8438*

0.6518 0.7332 0.7718 0.8829*

0.1684 0.2503 0.3452 0.4263

0.9867* 0.4889 0.1286 0.178

The experiment with 4-tert-octylphenol was not included in the analysis because data for the second sampling day was not available for stereological parameters. Asterisks indicate statistically significant correlations ( p < 0.05).

Table 2 Vitellogenin levels (measured in lg/mg prot) in adult male zebrafish treated with dibutylphthalate (DBP), methoxychlor (MXC), 4-tert-octylphenol (OP), 17a-ethynylestradiol (EE2), 17b-estradiol (E2) and the solvent control dimethylformamide (DMF) at each exposure time Treatment groups

Exposure time

DMF DBP DMF MXC DMF OP DMF EE2 DMF E2

7 days

15 days

0.018 T 0.004 0.017 T 0.007 (0) 0.014 T 0.004 2.43 T 0.22* (173) 0.03 T 0.003 4.55 T 1.8* (157) 0.016 T 0.003 12.17 T 1.83* (760) 0.017 T 0.002 10.36 T 4.1* (609)

0.017 T 0.004 0.022 T 0.005 (1.3) 0.017 T 0.006 6.38 T 3.47* (375) 0.032 T 0.002 11.14 T 1.85* (348) 0.021 T 0.004 26.72 T 5.17* (1272) 0.018 T 0.003 24.31 T 2.11* (1350)

Asterisks indicate statistically significant differences ( p < 0.05) with respect to controls of the corresponding experiment. Fold increment over values measured in control fish are given in parenthesis.

All treatments except DBP caused significant induction of vitellogenin (Vtg) at both sampling times (Table 2). Highest induction levels were found in E2 and EE2 treatment groups, followed by OP and MXC treatment groups. Induction was dependent on the exposure time, since values obtained at 15 days of exposure were approximately twice the values at 7 days. DBP did not alter Vtg levels in exposed male zebrafish. Correlation coefficients between Vtg levels and AOX activities obtained for the whole set of data and for each experiment are shown in Table 3. Results demonstrate a significant correlation between Vtg and AOX for the data as a whole and for MXC, OP and EE2 experiments. No gross differences were detected between gonads of fish from different treatments and those of control groups (Fig. 3a,b). Cysts in all developmental stages were observed and spermatogenic cells appeared normal. More subtle effects on spermatogenic cells were analyzed by immunohistochemistry of PCNA, an essential molecule for DNA

Table 3 Correlation indices between peroxisome proliferation marker enzyme acylCoA oxidase (AOX) activities and vitellogenin levels in different experiments where adult male zebrafish were treated with dibutylphthalate (DBP), methoxychlor (MXC), octylphenol (OP), 17a-ethynylestradiol (EE2) and 17b-estradiol (E2) TOTAL

DBP

MXC

OP

EE2

E2

Correlation 0.402** 0.217 0.720** 0.629* 0.699* 0.378 coefficient Signi. 0.001 0.499 0.008 0.028 0.011 0.226 N 60 12 12 12 12 12 TOTAL indicates all data from all experiments. One asterisk indicates statistically significant correlations at p < 0.05 and two asterisks at p < 0.01.

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Fig. 3. Micrographs of paraffin sections through the gonad of male zebrafish from the control and E2 treatment groups at 15 days of exposure. (a and b) 7Am sections stained with hematoxylin/eosin in control zebrafish (a) and in zebrafish exposed to E2 for 15 days (b). Spermatogenic cells at different developmental stages appear normal in both control and E2 groups. (c and d) 3 Am sections labeled immunohistochemically with PCNA in control zebrafish (c) and in zebrafish exposed to E2 (d) for 15 days. Nuclei of spermatogonial cells (arrows) are intensely labeled with PCNA while nuclei of spermatocytes (arrowheads) are labeled to a lesser extent.

Fig. 4. Surface density of PCNA-labeled cells in the gonad of adult male zebrafish treated for 7 and 15 days with dimethylformamide used as solvent (solid bars) and with dibutylphthalate (DBP), methoxychlor (MXC), 4-tert-octylphenol (OP), 17a-ethynylestradiol (EE2) and 17b-estradiol (E2) (open bars). Data are represented as mean T standard deviations. Asterisks indicate statistically significant differences ( p < 0.05) between solvent control and exposed groups within each exposure duration.

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Fig. 5. Number of PCNA-labeled spermatogonia and spermatocytes per unit area of gonad in adult male zebrafish treated for 7 and 15 days with dimethylformamide used as solvent (solid bars) and with 17a-ethynylestradiol (EE2) and 17b-estradiol (E2) (open bars). Data are represented as mean T standard deviations. Asterisks indicate statistically significant differences ( p < 0.05) between solvent control and exposed groups within each exposure duration.

synthesis during the S phase of the cell cycle and commonly used as cell proliferation marker. In the testis of adult male zebrafish, nuclei of spermatogonia were intensely labeled with PCNA while nuclei of spermatocytes were labeled to a lesser extent (Fig. 3c,d). This is possibly due to the fact that spermatogonial cells are the only cells undergoing S phase during meiosis and thus, PCNA labeling is gradually lost as spermatogenesis proceeds. No differences in the surface density of PCNA-labeled nuclei were observed in zebrafish exposed to DBP, MXC or OP but this parameter was significantly increased in zebrafish exposed to E2 and EE2 for 15 days (Figs. 3d and 4). This increase was due to the presence of a higher number of PCNA-immunolabeled spermatogonia per unit area of gonad while the number of labeled spermatocytes did not differ between control and exposed zebrafish (Fig. 5).

4. Discussion In the present work, we investigated if dibutylphthalate (DBP), methoxychlor (MXC), 4-tert-octylphenol (OP), 17aethynylestradiol (EE2) and 17b-estradiol (E2) were able to cause liver peroxisome proliferation in adult male zebrafish. Results indicated that all five compounds caused significant proliferation of liver peroxisomes. In addition, some of the compounds produced alterations in vitellogenin synthesis and spermatogenic cell proliferation, as discussed below separately for each compound. Phthalates are the most abundant man-made chemicals in the environment; they are produced in large quantities and used mainly to impart flexibility in plastics (Tyler et al., 1998). The ubiquity of phthalates in the aquatic environ-

ment is well known and their presence has been reported in rivers, waste and drinking waters, as well as in the marine environment, both in sediments and fish. In the present work, zebrafish exposed to 500 Ag/L of DBP during 7 or 15 days showed clear signs of peroxisome proliferation such as increased surface and numerical densities of peroxisomes and induced levels of peroxisomal acyl-CoA oxidase (AOX) activity. This agrees well with previous studies using phthalates in rodents (Reddy and Mannaerts, 1994) and mussels (Cancio et al., 1998; Orbea et al., 2002) and demonstrates for the first time that phthalates induce peroxisome proliferation in fish. Some phthalates, such as DBP and butyl benzyl phthalate, exhibit a weak positive result with in vitro estrogenicity tests (Jobling et al., 1995; Cadogan, 1999; Goksøyr et al., 2003). However, in the present work, there was no evidence indicating that DBP exposure produced estrogenic effects in adult male zebrafish. This agrees well with the results of other in vivo experiments with phthalates in mammals (Cadogan, 1999) and in fish such as rainbow trout (Knudsen et al., 1998). Organochlorine pesticides such as MXC are used to control a wide range of insect pests in field crops. MXC is considered a pro-estrogen in fish since it must undergo demethylation to the mono-or bis-demethylated derivatives in order to cause the estrogenic response (Schlenk et al., 1997, 1998; Miyashita et al., 2004). MXC metabolites reduce the binding affinity of E2 to the ER in fish in vivo, possibly binding to sites on the receptor other than the hormone binding domain (Nimrod and Benson, 1997). This binding of MXC derivatives to the ER appears to activate the receptor as our results show that MXC exposure caused induction of vitellogenin synthesis in adult male zebrafish. Other studies have demonstrated induction of vitellogenin in

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adult male fish exposed to MXC (Thorpe et al., 2000; Ankley et al., 2001; Hemmer et al., 2001). Additionally, our results suggest that MXC (or its derivatives) can interact with PPAR and/or PPRE in zebrafish since MXC exposure provoked increased peroxisomal density values and induced activity of AOX. A modest peroxisomal proliferation has previously been observed in liver of 4 –6 day old larvae of zebrafish after injection of 4-chloroaniline (Oulmi and Braunbeck, 1996) but not in adult female zebrafish exposed to the same compound (Braunbeck et al., 1990). Peroxisome proliferation has been reported in different fish species exposed to other organochlorine pesticides such as endosulfan or dieldrin (Arnold et al., 1995; Pedrajas et al., 1996). The chlorinated herbicide 2,4-dichlorophenoxyacetic acid has been demonstrated to induce peroxisome proliferation in Fundulus heteroclitus (Ackers et al., 2000). It is worth mentioning that of the five compounds tested, MXC was the only one causing significant induction of peroxisomal catalase activity. The expression of catalase is not regulated by PPARs and, indeed, catalase activity is only slightly induced or even inhibited by peroxisome proliferators (Reddy and Mannaerts, 1994; Cancio et al., 1998; Cajaraville et al., 2003a,b). In rats, no correlation between the degree of proliferation of peroxisomes and peroxisomal catalase activity has been found (Mehrotra et al., 1997). The increased activity of catalase in MXC-treated zebrafish could be related to increased levels of reactive oxygen species possibly produced during P450-mediated MXC metabolism (Bulger and Kupfer, 1989). Induction of cytochrome P450 system enzymes such as CYP2 isozymes, has been demonstrated in fish exposed to peroxisome proliferators (Haash et al., 1998). Alkylphenolic compounds are formed by microbial degradation of alkylphenol ethoxylates (APEs). APEs are non-ionic surfactants used in a variety of products including cleaning agents, textiles, agricultural chemicals, plastics, paper products, household cleaning agents, and personal care products. They are produced during sewage treatment, and have been detected in sewage effluents at concentrations up to 600 Ag/L (Tyler et al., 1998; Sole´ et al., 2000a). Alkylphenolic compounds show estrogenic activity in fish both in vitro and in vivo (Tyler et al., 1998). There is evidence that the estrogenic activity of alkylphenols is mediated by binding to the ER (Yadetie et al., 1999). Our results with vitellogenin induction in zebrafish exposed to OP (500 Ag/L) support this notion. Furthermore, as with DBP and MXC, OP exposure also caused increased surface and numerical densities of liver peroxisomes and induction of AOX activity, indicating interaction of OP (or its derivatives) with PPAR and/or PPRE. To the best of our knowledge this is the first report demonstrating that alkylphenols produce peroxisome proliferation in fish. The natural estrogenic steroid E2 and its synthetic analogues are also of potential concern to wildlife (Tyler et al., 1998). E2 and EE2 are present in sewage effluents, reservoirs, rivers and potable water at concentrations

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ranging from 0.2 to 20 ng/L (Aherne and Briggs, 1989; Tyler et al., 1998; Huang and Sedlak, 2001) but the effects of these coumpounds on fish have been studied using a wide range of concentrations from ng/L up to Ag/L (Veranicˆ and Pipan, 1992; Islinger et al., 2003) and even mg/L (Krisfalusi et al., 1998). In the present work, the selected exposure doses of 10 Ag/L of EE2 and E2 produced early sublethal effects such as liver peroxisome proliferation and induction of AOX activity and vitellogenin synthesis. Similarly, in adult male zebrafish exposed to 1 Ag/L E2 for 40 days, stimulation of vitellogenin synthesis was accompanied by an increase in the number and a decrease in the average diameter of peroxisomes causing a significant increase in the surface density of this organelle (Veranicˆ and Pipan, 1992). Adult fathead minnows (Pimephales promelas) exposed to EE2 for 3 weeks also showed increased plasma vitellogenin levels and higher peroxisome numbers in hepatocytes (Pawlowski et al., 2004). Conversely, exposure to EE2 and E2 caused a decrease in catalase activity of male zebrafish, in agreement with results of Sole´ et al. (2000b) in EE2-injected carp. This again reinforces the idea that catalase is not a good marker of peroxisome proliferation (Cajaraville et al., 2003a) as peroxisome proliferators generally cause inhibition of antioxidant enzymes (Cajaraville et al., 2003b). Although the five compounds tested did not produce gross alterations in gonad histology, immunohistochemistry of the cell proliferation marker PCNA showed that exposure to EE2 and E2 led to increased surface density of PCNAlabeled nuclei in the gonad. This increase was due to the presence of a higher number of PCNA-immunolabeled spermatogonia per unit area of gonad, suggesting a lower cell division rate of spermatogonia leading to a delay or even inhibition of spermatogenesis. Similarly, E2 and EE2exposed fish Sparus aurata showed inhibition of testicular growth and of male germ cell development beyond the spermatogonia stage, including mitosis (Condec¸ a and Canario, 1999). Adult active spermatogenic rainbow trout exposed to E2 also showed total inhibition of gonadal development (Billard et al., 1981). Thus, the technique of PCNA localization in spermatogenic cells offers potential as a novel biomarker of estrogenicity, able to determine subtle changes in spermatogenic cell proliferation. Overall, our results indicate that all five compounds tested produced significant peroxisome proliferation in liver of exposed adult male zebrafish in terms of increased peroxisomal volume, surface and numerical densities. These effects were more pronounced in the case of DBP and MXC exposed fish, but peroxisome proliferation was also evident in OP, EE2 and E2 exposed groups. Changes in peroxisomal density parameters were accompanied by significant induction of AOX activity, induction being more marked in the case of DBP. Further statistical analysis demonstrated the positive correlation between peroxisomal density parameters and AOX activity, as reported previously in mussels (Cancio et al., 1999). The finding that the five compounds

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tested produced peroxisome proliferation in zebrafish suggests a possible interaction of estrogens and xenoestrogens with PPARs and/or PPREs, as demonstrated recently using zebrafish hepatocytes primary cultures (Ibabe et al., in press). It is also possible that these compounds activate PPAR target genes (e.g., AOX) via ERE (Djouadi et al., 1998). Interestingly, 17b-hydroxysteroid dehydrogenase type IV, the key enzyme for E2 inactivation, has been shown to be induced by peroxisome proliferators through activation of PPARa (Fan et al., 1998). This could explain the results of the present work where the most potent peroxisome proliferator and AOX inducer DBP did not cause any effects on vitellogenin synthesis or other estrogenic effects. Conversely, E2 and EE2 treatments showed the less marked peroxisome proliferatory response while being the most potent inducers of vitellogenin synthesis and the only treatments causing alterations in spermatogenic cell proliferation. In conclusion, for the xenoestrogenic compounds MXC, OP and EE2, the significant correlation found between vitellogenin levels and AOX activity indicates an association between early estrogenic effects and liver peroxisome proliferation. No such an association occurs for typical peroxisome proliferators such as DBP. Further studies are needed at the molecular level to decipher the signal transduction pathways and molecules involved in the interaction between the two phenomena. The fact that xenoestrogens produce peroxisome proliferation in fish uncovers a new potential risk for fish exposed to these compounds as sustained peroxisome proliferation is strongly correlated to development of hepatocarcinomas in rodents (Reddy and Lalwani, 1983; Cajaraville et al., 2003a). Acknowledgements This work was supported by the Spanish Ministry of Science and Technology, CICYT, through project AMB990324 and by the European Commission (Research Directorate General, Environment Programme-Marine Ecosystems) through the BEEP project ‘‘Biological Effects of Environmental Pollution in Marine Coastal Ecosystems’’ (contract EVK3-CT2000-00025). BEEP project is part of the EC IMPACTS cluster. M. Ortiz-Zarragoitia is recipient of a pre-doctoral fellowship from the University of the Basque Country. The grant for consolidated research groups of the University of the Basque Country is also greatly acknowledged. References Ackers, J.T., Johnston, M.F., Haascj, M.L., 2000. Immunodetection of hepatic peroxisomal PMP70 as an indicator of peroxisomal proliferation in the mummichog, Fundulus heteroclitus. Mar. Environ. Res. 50, 361 – 365.

Aebi, H., 1974. Catalase. In: Bergemeyer, H.U. (Ed.), Methods of Enzymatic Analysis. Academic Press, Deerfield Beach, Florida, pp. 671 – 684. Aherne, G.W., Briggs, R., 1989. The relevance of the presence of certain synthetic steroids in the aquatic environment. J. Pharm. Pharmacol. 41, 735 – 736. Ankley, G.T., Jensen, K.M., Kahl, M.D., Korte, J.J., Makynen, E.A., 2001. Description and evaluation of a short-term reproduction test with the fathead minnow (Pimephales promelas). Environ. Toxicol. Chem. 20, 1276 – 1290. Arnold, H., Pluta, H.-J., Braunbeck, T., 1998. Simultaneous exposure of fish to endosulfan and disulfoyon in vivo: ultrastructural, stereological and biochemical reactions in hepatocytes of male rainbow trout (Onchorhynchus mykiss). Aquat. Toxicol. 33, 17 – 43. Arukwe, A., Goksøyr, A., 2003. Eggshell and egg yolk proteins in fish: hepatic proteins for the next generation: oogenetic, population, and evolutionary implications of endocrine disruption. Comp. Hepatol. 2:4. Barton, H.A., Andersen, M.E., 1998. Endocrine active compounds: from biology to dose response assessment. Crit. Rev. Toxicol. 28, 363 – 423. Billard, R., Breton, B., Richard, M., 1981. On the inhibitory effects of some steroids on spermatogenesis in adult rainbow trout (Salmo gairdneri). Can. J. Zool. 59, 1479 – 1487. Braunbeck, T., Storch, V., Bresch, H., 1990. Species-specific reaction of liver ultrastructure in zebrafish (Brachydanio rerio) and trout (Salmo gairdneri) after prolonged exposure to 4-chloroaniline. Arch. Environ. Contam. Toxicol. 19, 405 – 418. Bulger, W.H., Kupfer, D., 1989. Characteristics of monooxygenasemediated covalent binding of methoxychlor in human and rat liver microsomes. Drug Metab. Dispos. 17, 487 – 494. Cadogan, D.F., 1999. Health and environmental effects of phthalate plasticisers for poly(vinylchloride) —an update. Plast. Rubber Compos. 28, 476 – 481. Cajaraville, M.P., Orbea, A., Marigo´mez, I., Cancio, I., 1997. Peroxisome proliferation in the digestive epithelium of mussels exposed to the water accomodated fraction of three oils. Comp. Biochem. Physiol. C 117, 233 – 242. Cajaraville, M.P., Bebianno, M.J., Blasco, J., Porte, C., Sarasquete, C., Viarengo, A., 2000. The use of biomarkers to assess the impact of pollution in coastal environments of the Iberian Peninsula: a practical approach. Sci. Total Environ. 247, 295 – 311. Cajaraville, M.P., Cancio, I., Ibabe, A., Orbea, A., 2003a. Peroxisome proliferation as a biomarker in environmental pollution assessment. Microsc. Res. Tech. 61, 191 – 202. Cajaraville, M.P., Hauser, L., Carvalho, G., Hylland, K., Olabarrieta, I., Lawrence, A.J., Lowe, D., Goksøyr, A., 2003b. Chapter 2: Genetic damage and the molecular/cellular response to pollution. In: Lawrence, A.J., Hemingway, K.L. (Eds.), Effects of Pollution on Fish. Blackwell Science Ltd., Oxford, pp. 14 – 82. Cancio, I., Cajaraville, M.P., 2000. Cell biology of peroxisomes and their characteristics in aquatic organisms. Int. Rev. Cyt. 199, 201 – 293. Cancio, I., Orbea, A., Vo¨lkl, A., Fahimi, H.D., Cajaraville, M.P., 1998. Induction of peroxisomal oxidases in mussels: comparison of effects of lubricant oil and benzo(a)pyrene with two typical peroxisome proliferators on peroxisome structure and function in Mytilus galloprovincialis. Toxicol. Appl. Pharmacol. 149, 64 – 72. Cancio, I., Ibabe, A., Cajaraville, M.P., 1999. Seasonal variation of peroxisomal enzyme activities and peroxisomal structure in mussels Mytilus galloprovincialis and its relationship with the lipid content. Comp. Biochem. Physiol. C 123, 135 – 144. Condec¸a, J.B., Canario, A.V.M., 1999. The effect of estrogen on the gonads and on in vitro conversion of androstenedione to testosterone, 11ketotestosterone, and estradiol-17b in Sparus aurata (Teleostei, Sparidae). Gen. Comp. Endocrinol. 116, 59 – 72. Crews, D., Willingham, E., Skipper, J.K., 2000. Endocrine disruptors: present issues, future directions. Q. Rev. Biol. 75, 243 – 260. Djouadi, F., Weinheimer, C.J., Saffitz, J.E., Pitchford, C., Bastin, J., Gonza´lez, F.J., Kelly, D.P., 1998. A gender-related defect in lipid

M. Ortiz-Zarragoitia, M.P. Cajaraville / Comparative Biochemistry and Physiology, Part C 141 (2005) 133 – 144 metabolism and glucose homeostasis in peroxisome proliferatoractivated receptor a-deficient mice. J. Clin. Invest. 102, 1083 – 1091. Fan, L.Q., Cattley, R.C., Corton, J.C., 1998. Tissue-specific induction of 17b-hydroxysteroid dehydrogenase type IV by peroxisome proliferator chemicals is dependent on the peroxisome proliferator-activated receptor a. J. Endocrinol. 158, 237 – 246. Fatoki, O.S., Ogunfowokan, A.O., 1993. Determination of phthalate ester plasticizers in the aquatic environment of southwestern Nigeria. Environ. Int. 19, 619 – 623. Gillesby, B.E., Zacharewski, T.R., 1998. Exoestrogens: mechanisms of action and strategies for identification and assessment. Environ. Toxicol. Chem. 17, 3 – 14. Goksøyr, A., Arukwe, A., Larsson, J., Cajaraville, M.P., Hauser, L., Nilsen, B.M., Lowe, D., Matthiessen, P., 2003. Chapter 3: Molecular/cellular processes and the impact on reproduction. In: Lawrence, A.J., Hemingway, K.L. (Eds.), Effects of Pollution on Fish. Blackwell Science Ltd., Oxford, pp. 179 – 220. Haash, M.L., Henderson, M.C., Buhler, D.R., 1998. Induction of CYP2M1 and CYP2K1 lauric acid hydroxylase activities by peroxisome proliferating agents in certain fish species: possible implications. Mar. Environ. Res. 46, 37 – 40. Harris, C.A., Henttu, P., Parker, M.G., Sumpter, J.P., 1997. The estrogenic activity of phthalate esters in vitro. Environ. Health Perspect. 105, 802 – 811. Hemmer, M.J., Hemmer, B.L., Bowman, C.J., Kroll, K.J., Folmar, L.C., Marcovich, D., Hoglund, M.D., Denslow, N.D., 2001. Effects of pnonylphenol, methoxychlor, and endosulfan on vitellogenin induction and expression in sheepshead minnow (Cyprinodon variegatus). Environ. Toxicol. Chem. 20, 336 – 343. Huang, C.-H., Sedlak, D.L., 2001. Analysis of estrogenic hormones in municipal wastewater effluent and surface water using enzyme-linked immunosorbent assay and gas chromatography tandem mass spectrometry. Environ. Toxicol. Chem. 20, 33 – 139. Ibabe, A., Grabenbauer, M., Baumgart, E., Fahimi, H.D., Cajaraville, M.P., 2002. Expression of peroxisome proliferator-activated receptors in zebrafish (Danio rerio). Histochem. Cell Biol. 118, 231 – 239. Ibabe, A., Herrero, A., Cajaraville, M.P., in press. Modulation of peroxisome proliferator-activated receptors (PPARs) by PPARa-and PPARc-specific ligands and by 17b-estradiol in isolated zebrafish hepatocytes. Toxicol. In Vitro. Islinger, M., Willimski, D., Vo¨lkl, A., Braunbeck, T., 2003. Effects of 17aethinylestradiol on the expression of three estrogen-responsive genes and cellular ultrastructure of liver and testes in male zebrafish. Aquat. Toxicol. 62, 85 – 103. Jobling, S., Reynolds, T., White, R., Parker, M.G., Sumpter, J.P., 1995. A variety of environmentally persistent chemicals, including some phthalate plasticizers, are weakly estrogenic. Environ. Health Perspect. 103, 582 – 587. Kashiwada, S., Ohnishi, Y., Ishikawa, H., Miyamoto, N., Magara, Y., 2001. Comprehensive risk assessment of estradiol-17b, p-nonylphenol, and bisphenol-A in river water in Japan. Environ. Sci. 8, 89 – 102. Keller, J.M., Collet, P., Bianchi, A., Huin, C., Bouillaud-Kremarik, P., Becuwe, P., Schohn, H., Domenjoud, L., Dauc¸a, M., 2000. Implications of peroxisome proliferator-activated receptors (PPARs) in development, cell life status and disease. Int. J. Dev. Biol. 44, 429 – 442. Kirk, C.J., Bottomley, L., Minican, N., Carpenter, H., Shaw, S., Kohli, N., Winter, M., Taylor, E.W., Waring, R.H., Michelangeli, F., Harris, R.M., 2003. Environmental endocrine disrupters dysregulate estrogen metabolism and Ca2+ homeostasis in fish and mammals via receptorindependent mechanisms. Comp. Biochem. Physiol., Part A 135, 1 – 8. Knudsen, F.R., Arukwe, A., Pottinger, T.G., 1998. The in vivo effect of combinations of octylphenol, butylbenzylphthalate and estradiol on liver estradiol receptor modulation and induction of zona radiata proteins in rainbow trout: no evidence of synergy. Environ. Pollut. 103, 75 – 80. Krisfalusi, M., Eroschenko, V.P., Cloud, J.G., 1998. Methoxychlor and estradiol-17b affect alevin rainbow trout (Oncorhynchus mykiss)

143

mortality, growth, and pigmentation. Bull. Environ. Contam. Toxicol. 61, 519 – 526. Lemberger, T., Braissant, O., Juge-Aubry, C., Keller, H., Saladin, R., Staels, B., Auwerx, J., Burger, A., Meier, C.A., Wahli, W., 1996. PPAR tissue distribution and interactions with other hormone-signaling pathways. Ann. N.Y. Acad. Sci. 804, 231 – 251. Ma, H., Sprecher, H.W., Kolattukudy, P.E., 1998. Estrogen-induced production of a peroxisome proliferator-activated receptor (PPAR) ligand in a PPARc-expressing tissue. J. Biol. Chem. 273, 30131 – 30138. Magalhaˆes, M.M., Magalhaˆes, M.C., 1997. Peroxisomes in adrenal steroidogenesis. Microsc. Res. Tech. 36, 493 – 502. Markus, M., Husen, B., Adamski, J., 1995. The subcellular localization of 17b-hydroxysteroid dehydrogenase type 4 and its interaction with actin. J. Steroid Biochem. Mol. Biol. 55, 617 – 621. Matthiessen, P., Sumpter, J.P., 1998. Effects of estrogenic substances in the aquatic environment. In: Braunbeck, T., Hinton, D.E., Streit, B. (Eds.), Fish Ecotoxicology. Birkha¨user Verlag, Basel, Switzerland, pp. 319 – 335. Mehrotra, K., Morgenstern, R., Lundquist, G., Becedas, L., Ahlberg, M.B., Georgellis, A., 1997. Effects of peroxisome proliferators and/or hypothyroidism on xenobiotic-metabolizing enzymes in rat testis. Chem.Biol. Interact. 104, 131 – 145. Miyashita, M., Shimada, T., Nakagami, S., Kurihara, N., Miyagama, H., Akamatsu, M., 2004. Enantioselective recognition of mono-demethylated methoxychlor metabolites by the estrogen receptor. Chemosphere 54, 1273 – 1276. Nimrod, A.C., Benson, W.H., 1997. Xenobiotic interaction with and alterations of channel catfish estrogen receptor. Toxicol. Appl. Pharmacol. 147, 381 – 390. Orbea, A., Ortiz-Zarragotia, M., Cajaraville, M.P., 2002. Interactive effects of benzo(a)pyrene and cadmium and effects of di(2-ethylhexyl)phthalate on antioxidant and peroxisomal enzymes and peroxisomal volume density in the digestive gland of mussels Mytilus galloprovincialis Lmk. Biomarkers 7, 33 – 48. ¨ rn, S., Holbech, H., Madsen, T.H., Norrgren, L., Petersen, G.I., 2003. O Gonad development and vitellogenin production in zebrafish (Danio rerio) exposed to ethinylestradiol and methyltestosterone. Aquat. Toxicol. 65, 397 – 411. Oulmi, Y., Braunbeck, T., 1996. Toxicity of 4-chloroaniline in early lifestages of zebrafish (Brachydanio rerio): I. Cytopathology of liver and kidney after microinjection. Arch. Environ. Contam. Toxicol. 30, 390 – 402. Pawlowski, S., van Aerle, R., Tyler, C.R., Braunbeck, T., 2004. Effects of 17a-ethinylestradiol in a fathead minnow (Pimephales promelas) gonadal recrudescence assay. Ecotoxicol. Environ. Saf. 57, 330 – 345. Pedrajas, J.R., Lo´pez-Barea, J., Peinado, J., 1996. Dieldrin induces peroxisomal enzymes in fish (Sparus aurata) liver. Comp. Biochem. Physiol., C 115, 125 – 131. Reddy, J.K., Lalwani, N.D., 1983. Carcinogenesis by hepatic peroxisome proliferators: evaluation of the risk of hypolipidemic drugs and industrial plasticizers to humans. CRC Crit. Rev. Toxicol. 12, 1 – 58. Reddy, J.K., Mannaerts, G.P., 1994. Peroxisomal lipid metabolism. Annu. Rev. Nutr. 14, 343 – 370. Small, G.M., Burdett, K., Connock, M.J., 1985. A sensitive spectrophotometric assay for peroxisomal acyl-CoA oxidase. Biochem. J. 227, 205 – 210. Schlenk, D., Stresser, D.M., McCants, J.C., Nimrod, A.C., Benson, W.H., 1997. Influence of b-naphthoflavone and methoxychlor pretreatment on the biotransformation and estrogenic activity of methoxychlor in channel catfish (Ictalurus punctatus). Toxicol. Appl. Pharmacol. 145, 349 – 356. Schlenk, D., Stresser, D.M., Rimoldi, J., Arcand, L., McCants, J., Nimrod, A.C., Benson, W.H., 1998. Biotransformation and estrogenic activity of methoxychlor and its metabolites in channel catfish (Ictalurus punctatus). Mar. Environ. Res. 46, 159 – 162. Segner, H., Caroll, K., Fenske, M., Janssen, C.R., Maack, G., Pascoe, D., Scha¨ffers, C., Vanderbergh, G.F., Watts, M., Wenzel, A., 2003.

144

M. Ortiz-Zarragoitia, M.P. Cajaraville / Comparative Biochemistry and Physiology, Part C 141 (2005) 133 – 144

Identification of endocrine-disrupting effects in aquatic vertebrates and invertebrates: report from the European IDEA project. Ecotoxicol. Environ. Saf. 54, 302 – 314. Sole´, M., Lo´pez de Alda, M.J., Castillo, M., Porte, C., Ladegaard, K., Barcelo´, D., 2000a. Estrogenicity determination in sewage treatment plants and surface waters from the Catalonian area (NE Spain). Environ. Sci. Technol. 34, 5076 – 5083. Sole´, M., Porte, C., Barcelo´, D., 2000b. Vitellogenin induction and other biochemical responses in carp, Cyprinus carpio, after experimental injection with 17a-ethynylestradiol. Arch. Environ. Contam. Toxicol. 38, 494 – 500. Thorpe, K.L., Hutchinson, T.H., Hetheridge, M.J., Sumpter, J.P., Tyler, C.R., 2000. Development of an in vivo screening assay for estrogenic chemicals using juvenile rainbow trout (Oncorhynchus mykiss). Environ. Toxicol. Chem. 19, 2812 – 2820. Tyler, C.R., Jobling, S., Sumpter, J.P., 1998. Endocrine disruption in wildlife: a critical review of the evidence. Crit. Rev. Toxicol. 28, 319 – 361.

Veranicˆ, P., Pipan, N., 1992. The relationship between endoplasmatic reticulum and peroxisomes in fish hepatocytes during estradiol stimulation and after cessation of vitellogenesis. Period. Biol. 94, 29 – 34. Wang, X., Kilgore, M.W., 2002. Signal cross-talk between estrogen receptor alpha and beta and the peroxisome proliferator-activated receptor gamma1 in MDA-MB-231 and MCF-7 breast cancer cells. Mol. Cell. Endocrinol. 194, 123 – 133. WHO/IPCS World Health Organization/International Petroleum Chemical Safety, 2002. Global assessment of the state-of-the-science of endocrine disruptors. In: Damstra, T., Barlow, S., Bergman, A., Kavlock, R., Van der Kraak, G. (Eds.), WHO/PCS/EDC/02.2. World Health Organization, Geneva, Switzerland. Available from: http://ehp.niehs.nih.gov/ who/. Yadetie, F., Arukwe, A., Goksøyr, A., Male, R., 1999. Induction of hepatic estrogen receptor in juvenile Atlantic salmon in vivo by the environmental estrogen, 4-nonylphenol. Sci. Total Environ. 233, 201 – 210.