General and Comparative Endocrinology xxx (2014) xxx–xxx

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Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response I. Traversi a,c,⇑, G. Gioacchini b, A. Scorolli b, D.G. Mita a, O. Carnevali b,c, A. Mandich a,c a

Department of Earth, Environment and Life Sciences (DiSTAV), University of Genoa, Italy Department of Environment and Life Sciences, University of Marche, Ancona, Italy c Interuniversity Consortium INBB, Rome, Italy b

a r t i c l e

i n f o

Article history: Available online xxxx Keywords: Sea bream juveniles Nonylphenol 4-Tert-octylphenol Morphological endpoints Biomarkers

a b s t r a c t A wide range of endocrine disrupter chemicals can mimic steroid hormones causing adverse health effects. Nonylphenol (NP) and t-octhylphenol (t-OP) are man-made alkylphenolic environmental contaminants possessing controversial endocrine disruption properties. This study has investigated the effects of NP and t-OP enriched diets on hepatic tissue and biotransformation activities in the liver. To this aim, sea bream juveniles were fed with commercial diet enriched with three different doses of NP (NP1: 5 mg/kg bw, NP2: 50 mg/kg bw and NP3: 100 mg/kg bw) or t-OP (t-OP1: 5 mg/kg bw, t-OP2: 50 mg/kg bw and t-OP3: 100 mg/kg bw) for 21 days. A significant increase of the hepatosomatic index was observed in NP1 and t-OP1. Alteration of liver morphology was observed in both NP and t-OP exposed juveniles although the most altered endpoints were observed in t-OP2 with 100% of tissue degeneration. Ethoxyresorufin-O-deethylase activity was significantly inhibited by NP and t-OP (p < 0.05), while catalase activity was significantly induced, at both doses. A different pattern of protein expression of different isoforms of both vitellogenin and zona radiata protein was evidenced within the treatments. In addition, a significant increase in the abundance of the stress induced heat shock protein 70 gene in the liver of t-OP2 fish and a significant increase in the abundance of the estrogen induced cathepsin D gene in the liver of NP1 and t-OP2 fish, were observed. Finally, estradiol-17b (E2) and testosterone (T) plasma levels and E2/T showed significantly different patterns in NP and t-OP exposed against control fish. Ó 2014 Elsevier Inc. All rights reserved.

1. Introduction In the last years many concerns have been raised regarding to alkylphenols (APs), since they represent one of the most important categories of Endocrine Disrupting Chemicals (EDCs). They are the degradation products of alkylphenol polyethoxylates (APEOs), an important group of non-ionic surfactants commonly used in many formulated products for industrial, agricultural, and domestic applications (Du et al., 2008). About 60% of APEOs end up in the aquatic environment: they are incompletely degraded to alkylphenols (APs), such as nonylphenol (NP) and 4-tert-octylphenol (t-OP), stable hydrophobic substances that tend to bioaccumulate in tissues of aquatic organisms (Isidori et al., 2006). The interest in their

⇑ Corresponding author at: Dipartimento di Scienze della Terra, dell’Ambiente e della Vita (DiSTAV), Università degli Studi di Genova, Italia, Corso Europa 26, 16132 Genova, Italia. Fax: +39 010 3538047. E-mail address: [email protected] (I. Traversi).

fate and environmental distribution has increased as consequence of their extensive use and applications and their potential effects as EDCs (Bouzas et al., 2011). The systematic study of APs toxicity on fish can not only be of significance for the protection of fishery resources, but also provide evidence of the potential effects of APs on human health via the food chain (Du et al., 2008). The impacts of APs in the environment include estrogenic effects in aquatic organisms (Del Giudice et al., 2012; Zha et al., 2007) as well in reptiles (Verderame et al., 2011), birds (Oshima et al., 2012; Roig et al., 2014) and mammals (Zhang et al., 2014). The interference on fish reproduction by NP is well known: it is able to activate fish estrogen-dependent gene expression and elevate the plasma vitellogenin (VG) and zona radiata proteins (ZRP) in both males and females (Ackermann et al., 2002; Maradonna and Carnevali, 2007), to induce intersexuality (Jobling et al., 1996), to inhibit spermatogenesis and alter gonadosomatic index (Sepùlveda et al., 2003), above interfering with biotransformation activity (Teles et al., 2005). More recently it has been demonstrated that also t-OP may be able to promote the ZRP synthesis in males of

http://dx.doi.org/10.1016/j.ygcen.2014.06.015 0016-6480/Ó 2014 Elsevier Inc. All rights reserved.

Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015

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I. Traversi et al. / General and Comparative Endocrinology xxx (2014) xxx–xxx

Cichlosoma dimerus (Genovese et al., 2011), above multiple adverse effects such as increase of hepatic Glutathione-S-Transferase (GST) activity, decrease of plasma testosterone and cortisol levels, histopathological changes in liver, spleen, intestinal tract and testis (Carrera et al., 2007; Du et al., 2008). Since biotransformation of chemicals is a requisite for the detoxification and the subsequent excretion (Gravato and Santos, 2003), phase I and phase II biotransformation activities are currently used in environmental risk assessment like biomarker of exposure to many xenobiotics, among different EDCs. In addition heat shock protein 70 (HSP70) and cathepsin D (CATD) have recently been introduced as bioindicators of endocrine disruption (Carnevali and Maradonna, 2003). HSP70 belongs to a family of proteins which are expressed in response to a wide variety of stressors protecting cells by binding and refolding damaged proteins (Georgopoulous and Welch, 1993; Iwama et al., 1998) and numerous studies have clearly demonstrated that EDCs induce the expression of HSP70 in fish (Maradonna and Carnevali, 2007). CATD is an aspartic protease enzyme involved in various physiological pathways, including intracellular proteolysis (Barret, 1977) and its regulation by estrogenic EDCs (Maradonna and Carnevali, 2007) is well supported by the presence of a documented ERE in the promoter region (Augereau et al., 1994). The severe and/or prolonged physiology and biochemical variations in organs due to chemicals can lead to structural alterations. Thus, histology analysis represents a useful tool to study toxic effects on organisms under mild or long time exposure to xenobiotics (van Dyk et al., 2009). Nevertheless, the absence of numerical data in classical qualitative histo-pathological approaches makes it difficult to establish cause–effect relationships between pathology and contamination patterns. For this, recently different researchers focus their studies on histo-pathological indices to provide numerical data based on a semi-quantitative approach (Costa et al., 2009). Since the liver is the key organ for the defence against toxicants, through metabolism and the excretion of toxic substances, histological and biochemical changes in hepatic tissue are commonly used as biomarker of exposure to environmental pollutants (Hinton et al., 2001) The aim of this study was to provide important insights with respect to possible adverse effects on reproduction, liver toxicity and status stress regarding the potential impacts on wild Seabream juveniles ingesting prey containing NP or t-OP or farmed juveniles who could be exposed to 4-NP in their diet formulation. The oral exposure to graded concentrations (5, 50 and 100 mg/kg bw) of NP and t-OP in the present investigation was chosen in order to simulate a natural route of exposure to xenoestrogens and therefore goes beyond a purely environmental aim. The endpoints investigated after three weeks of exposure under laboratory conditions were liver morphology and enzyme activities, plasma sex steroid levels and ratio in order to allow an objective comparison between exposed and control specimens. Moreover possible reproductive disruption in treated fish was investigated with specific biomarkers such as the occurrence of VG and ZRP in the blood of male fish. In addition gene expression levels of HSP70 and CATD were investigated as biomarker of stress status. Finally the correlation between the occurrence of estrogenic effects of environmental APs and stress bioindicator was analyzed using univariate and multivariate statistical analysis. The doses of chemicals added to the food were chosen according to previously published work Bjerregaard et al. (2007) and is quite difficult to compare the dose administered via the diet to the amount present in the environment. Anyway, in the work of Pickford et al. (2003), fish were exposed to NP for 2 weeks either via the water to relevant environmental concentrations: 1, 10 or 50 lg/l or 100, 500 or 1000 lg/day via the diet. A 10-fold greater

sensitivity for NP in fish exposed via the water compared with exposure via the oral route was observed. The quantities of NP chosen for our study, fall in the same range as those mentioned. 2. Materials and methods 2.1. Fish maintenance Gilthead seabream juveniles, Sparus aurata (10.6 ± 3.7 g initial weight), were obtained from the Italian fish farm Orbetello Pesca lagunare, Grosseto, Italy. Acclimatization and rearing was carried out in a closed system equipped with biological, mechanical and UV filtration under the following conditions: temperature 20 ± 1 °C, salinity 35 ± 1 ppt, oxygen 6 ± 1 mg/l and photoperiod 12hL:12hD. The water in the tanks was changed up to 10 times a day through a dripping system. Ammonia and nitrite were constantly kept below 0.01 mg L1. The fish, at the initial stocking density of 10 individuals/100-L were fed by hand daily, ad libitum, using a commercial diet for seabream from Nutreco (Norway), with increasing pellet diameter according to the weight of the fish. Once acclimated, fish were divided in 4 groups of 20 specimens, in duplicates, and fed considering that the amount of feed administered was equal to 1% of the body weight of the fish. Each tank was fed once a day for 21 days and the feed administered in such a way as to ensure to all individuals the same amount of food. 2.2. Preparation of food Food prepared according to Bjerregaard et al. (2007). Commercial seabream food was crushed in a mortar and mixed with water at a 2:3 ratio to obtain a food homogenate. NP and t-OP (Sigma Aldrich, St. Louis, MO) was dissolved in 1 ml of acetone and added to 100 ml of food homogenate while it was stirred. Only the solvent was added to the control food. The food homogenate was left for 1–2 h for the acetone to evaporate. Three grams of gelatine powder (Sigma) were mixed into the 100 ml of food homogenate. After gentle heating, the homogenate was dispensed on a plate, allowed to cool to 5 °C overnight, and cut into cubes and stored at 20 °C until used in experiments. The amounts of NP and t-OP added to the homogenates were adjusted to give the nominal dose according to the growth rate of the fish, they were weighted every seven days. The tanks, in duplicate were fed as follows: Control (Cr) fed on commercial pelleted food, NP1 fed on commercial food enriched with 5 mg/kg bw NP, NP2 fed on commercial food enriched with 50 mg/kg bw NP, NP3 fed on commercial food enriched with 100 mg/kg bw NP, t-OP1 fed on commercial food enriched with 5 mg/kg bw t-OP, t-OP2 fed on commercial food enriched with 50 mg/kg bw t-OP, t-OP3 fed on commercial food enriched with 100 mg/kg bw t-OP. After three weeks of exposure, ten juveniles from each tank were sacrificed by a lethal overdose of anaesthesia (500 mg/L MS-222 (3-aminobenzoic acid ethyl ester) buffered to pH 7.4; Sigma) and total length (TL) and body weight (BW) were measured. Blood was taken from the heart using chilled heparinized syringes, and centrifuged (1200g, 15 min). Plasma was deeply frozen for subsequent sex steroid measurements and Western Blot analysis. Livers from all the experimental groups were dissected, weighed for the calculation of hepatosomatic index (HSI) and three fragments were served for next histological procedures, mRNA extraction and enzymatic activities, respectively. The HSI was calculated as liver weight/total body weight 100%. Procedures were performed in accordance with the Guidelines on the Handling and Training of Laboratory Animals by the Universities Federation for Animal Welfare (UFAW, 1992) and with the Italian animal welfare legislation (D.L. 116/92) and were approved

Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015

I. Traversi et al. / General and Comparative Endocrinology xxx (2014) xxx–xxx

by the Ethics Committee of Università Politecnica delle Marche (Prot #24/INT/CESA12-16).

2.3. Liver histology Fragments of liver were fixed in paraformaldehyde (4%) for 24 h at room temperature and served in 70% ethylic alcohol. Bio-Plast (Bio-Optica) embedded sections, 5 lm thick, were stained with Hematoxylin–Eosin for general histological examination and Mann Dominici for melanomacrophage centres (MMCs) analysis. Sections were examined by a Leica DMRB, images were visualized through the Leica Camera Microsystem DFC 420C and acquired through the software Leica Application Suite 3.4.0. To assess the quality of the hepatic tissue, presence/absence of the following histo-pathological alterations was recorded and evaluated against the grading system from Richardson et al. (2010) adapted for this study. In particular, the grade assigned to each liver corresponds to the average percentage of the endpoint occurrence in 9 areas of 97.2 mm2 randomly chosen in 3 sections, separated by at least 50 lm each other. The following endpoints were considered: (1) lipid accumulation (vacuolation) within hepatocytes: mild, moderate and severe; (2) lost of the typical hepatic cord and hepatopancreatic structures, (3) ceroids, (4) hemorrhagies, (5) blood vessel congestion, (6) hydropic change, (7) melanomacrophage centres (MMc), (8) lymphocytes and (9) liver parenchyma degeneration.

2.4. Enzymatic activities Liver sub-samples, preserved in 80 °C, were used for measuring enzymatic activities: ethoxyresorufin-O-deethylase (EROD), glutathione S-transferase (GST) and catalase (Cat). Livers were homogenate in pH 7.4, 10 mM Tris 250 mM saccharose 1 mM Na2EDTA buffer and centrifuged at 500g for 10 min. The resulting supernatant was divided in different aliquots. Total proteins were measured by the Bradford method, using blue Coomassie reagent and bovine serum albumin (BSA) as standard (Bradford, 1976).

2.4.1. EROD activity Ethoxyresorufin-O-deethylase (EROD) activity was determined as described in Suteau et al. (1988). The samples were centrifuged at 9000g for 20 min, to obtain S9 fraction. Mix reaction containing 200 mM Na2HPO4 50 mM KH2PO4 pH 7.4, 0.25 mM NADP, 1.23 lM 7-ethoxyresorufin and 1 unit ml-1 G-6-PDH was incubated at 30 °C for 5 min and 1 ml of this was added to all S9 fractions. The reaction was stopped by adding 2 ml of ice-cold acetone, immediately in S9 fraction time zero and after 5 min in S9 fraction time 5. Samples were centrifuged at 9000g for 15 min and 7-hydroxyresorufin fluorescence was determined using a Perkin Elmer LS50B spectrofluorometer at 537/583 excitation/emission wavelengths. Activity was calculated as the amount of resorufin (pmol) generated per milligram of protein per minute of reaction time.

2.4.2. GST activity Glutathione S-transferase (GST) activity was measured in the cytosolic fraction of liver, using 1-chloro-2,4-dinitrobenzene (CDNB) as substrate (Booth et al., 1961). The final reaction mixture contained 100 mM Na2HPO4/NaH2PO4 pH 6.8, 1 mM CDNB e 1 mM reduced glutathione in a total volume of 1 ml. The change in absorbance was recorded at 340 nm and the enzyme activity was calculated as nmol CDNB conjugate formed min1 mg1 protein using a molar extinction coefficient of 9.6 mM1 cm1.

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2.4.3. Cat activity Catalase (Cat) activity was measured according to Bergmeyer et al. (1983). The assay was performed in final volume of 1 ml, containing 50 mM NaH2PO4/Na2HPO4 pH 6.8 and 50 mM 30% H2O2 as substrate. After the adding of sample the decrease in absorbance at 240 nm due to H2O2 (hydrogen peroxide) consumption is followed for 3 min. The CAT activity was determined as the difference in the absorbance per unit of time (e = 0.04 mM1 cm1), and expressed as lmol H2O2 consumed min1 mg1 of total protein concentration 2.5. RNA extraction, cDNA synthesis and Real time PCR Total RNA extraction from livers was performed using Minikit RNAeasyÒ (Qiagen) extraction kits, following the manufacturer’s protocol. Total RNA extracts were eluted in 30 ll of RNAse-free water. Final RNA concentrations were determined with a NanoDrop™ 1000 Spectrophotometer (Thermo Scientific) and the RNA integrity was verified by ethidium bromide staining of 28S and 18S ribosomal RNA bands on 1% agarose gel. RNA was stored at 80 °C until use. Total RNA was treated with DNAse (10 IU at 37 °C for 10 min, MBI Fermentas), a total amount of 1 lg of RNA was used for cDNA synthesis, employing iScript cDNA Synthesis Kit (Bio-Rad). PCRs were performed with the SYBR green method in an iQ5 iCycler thermal cycler (Bio-Rad,). Triplicate PCR reactions were carried out for each sample analyzed. The reactions were set on a 96well plate by mixing; each reaction mixture consisted of 1 ll of diluted (1/10) cDNA, 10 ll of 2 concentrated iQ TM SYBR Green Supermix (Bio-Rad), containing SYBR Green as a fluorescent intercalating agent, 0.3 lM of forward primer and 0.3 lM of reverse primer. The thermal profile for all reactions was 3 min at 95 °C and then 45 cycles of 20 s at 95 °C, 20 s at 60 °C and 20 s at 72 °C. Fluorescence monitoring occurred at the end of each cycle. Additional dissociation curve analysis was performed and showed in all cases one single peak. Beta actin (ACT) and Elongation factor 1 alpha (EF1A) were used as internal standards in each sample in order to standardize the results by eliminating variation in mRNA and cDNA quantity and quality. No amplification products were observed in negative controls and no primer–dimer formations were observed in the control templates. The data obtained were analyzed using the iQ5 optical system software version 2.0 (Bio-Rad) including GeneEx Macro iQ5 Conversion and genex Macro iQ5 files. Modification of gene expression is represented with respect to the control. The primer sequences used are reported in Table 1. 2.6. Western blotting For the VG and ZRP assays, plasma samples were electrophoresed and transferred to polyvinylidenedifluoride (PVDF) membranes (Maradonna and Carnevali, 2007). Briefly, 20 lg of each protein sample was separated using 4% stacking and 10% separating sodium dodecyl sulfate polyacrylamide gel electrophoresis (SDS– PAGE) (Laemmli, 1970), and electroblotted onto a filter using a mini trans-blot electrophoretic transfer cell (all from Bio-Rad). Transfer was carried out for 30 min using Bio-Rad’s Trans-BlotÒ Turbo™ Transfer System. The membrane was soaked in 5% Nonidet-P40 for 1 h to remove SDS and incubated with 2% bovine serum albumin (BSA; Sigma) in PBS. The S. aurata VG primary antibody (Mosconi et al., 1998) was diluted 1:10,000 in a solution containing 2% BSA, 0.01% NaN3 in PBS, incubated for 1 h at room temperature (about 30 °C), and rinsed 3 times with PBS plus 0.05% Tween 20. For the ZRP assay, the primary polyclonal antibody rabbit anti-salmon ZRP was purchased from Biosense Laboratories AS, Bergen, Norway and diluted 1:1000 in a solution containing

Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015

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I. Traversi et al. / General and Comparative Endocrinology xxx (2014) xxx–xxx Table 1 List of primers used for real time PCR analyses. Gene name

Forward primer

Reverse primer

Efficiency

Beta ACT EF1 alpha CAT D HSP70

GGTACCCATCTCCTGCTCCAA AGTCCACCTCCACCGGTCAT TCCGTTCACTGCTCCTTGTTAG ACGGAGAGTCGATTTCGATG

GAGCGTGGCTACTCCTTCACC AGGAGCCCTTGCCCATCTC AGACTGCCGCTTCCATACTG GAAGGACATCAGCGACAACA

105.10% 99.70% 95.70% 99.80%

2% BSA, 0.01% NaN3 in PBS, and incubated for 2 h at room temperature (about 20 °C) and rinsed three times with BSA plus 0.05% Tween 20. For both assays, the second antibody solution (HRP-conjugated anti-rabbit IgG; BioRad) diluted 1:1000 in 2% BSA in PBS buffer was incubated for 1 h. The filter was rinsed again with PBS without Tween 20. The blot has been developed using as substrate ECL/Plus Western Blotting Detection System (Amersham Biosciences, UK). Densitometric analysis was performed using ImageJ software for Windows. 2.7. Sex steroid immunoassays Plasma samples were extracted twice with diethyl ether, and concentrations of Estradiol-17b (E2) and Testosterone (T) were measured by commercially available competitive enzyme-linked immunosorbent assays (ELISA) (Cayman Chemicals, Ann Arbour, MI, USA) which measure the total amount of steroids in plasma. 2.8. Calculations and statistical analyses The hepatosomatic index (HSI) was calculated as (LW/BW) 100. Significant differences in body and liver weights, HIS, enzymatic activities, sex steroid measures, gene expression and protein levels between treatment groups were determined using a factorial analysis of variance (one-way ANOVA) followed by Bonferroni’s multiple comparison test, using the statistical software package Prism5 (Graphpad Software, Inc. USA). Significance was accepted at p < 0.05. Significant differences in morphological endpoint data were tested with non-parametric Kruskal–Wallis test and, if necessary, a pairwise comparison with a Mann–Whitney test with Bonferroni correction was applied. Alpha level was set at 0.05 (0.017 for post hoc comparison) and all the statistical analysis were performed using SPSS v.20 (IBM Corp.). Data are presented as mean ± S.D. in all experiments. At the end, data were submitted all together, to multivariate analysis in order to better understand the selectivity of each variable to discriminate between groups. Where significant differences between groups were detected through the PERMANOVA test, similarity percentage analysis (SIMPER), Bray Curtis dissimilarity, was used to identify the best set of variables that allow to distinguish between groups (Primer 6 & PERMANOVA+). 3. Results Exposures to NP3 and t-OP3 resulted lethal, showing a 50% mortality at the end of the second week, and were not considered. 3.1. Body and liver weights No significant differences were observed in mean BW and TL between control and exposed groups. Hepatosomatic index (HSI) was significantly higher in NP1 (1.57 ± 0.3) against control (0.82 ± 0.2) and NP2 (0.99 ± 0.3) groups. Similarly, HSI was significantly higher in t-OP1 (1.55 ± 0.3) against control (0.82 ± 0.2) and t-OP2 (0.69 ± 0.2).

3.2. Histological endpoints Controls showed the characteristic organization of the sea bream liver as hepatopancreatic structure with the exocrine pancreatic tissue developed around the hepatic portal vein (Fig. 1A). The hepatocytes showed a round central nucleus, densely stained, and one prominent nucleolus and were organized in cords with one or two cell thickness around blood sinusoids. The vascularisation was normal and the bilious ducts were pervious. Small and medium round vesicles were observed in the cytoplasm of hepatocytes with mild and moderate lipid accumulation (Fig. 1B and C), respectively. Severe lipid accumulation was described when vesicles fuse to form a single large vacuole forcing cytoplasm and nucleus at the periphery of the cell (Fig. 1D). In this latter, the nucleus was characterized by chromatin condensation and increased optical density. In general, the mild and moderate lipid vacuolation did not alter the tissue organization in control livers. Mild, moderate and severe lipid accumulations were observed in the hepatic tissue of almost all the groups with different percentage proportion (Table 2). Control livers were mainly characterised by moderate lipid accumulation. In NP1 exposed juveniles, moderate lipid accumulation started to decrease reaching a significantly lower percentage in NP2 exposed group. In parallel, a higher percentage of hepatic tissue was interested by severe lipid accumulation in both NP1 and NP2 exposed groups. In NP2 a significant percentage of liver was characterised by absence of lipid accumulation. On the contrary, livers of t-OP exposed sea bream juveniles were interested by the occurrence of lipid accumulation showing a significant increase of severe lipid accumulation in both t-OP1 and t-OP2 exposed groups. None of the considered endpoint was observed in control juveniles. Hyalinization and lymphocyte infiltration were never observed even in livers of exposed fish. Significant loss of cord structure was observed in 18.52 ± 4.99% and 44.46 ± 8.55% NP1 and NP2 groups, respectively. The increasing occurrence of this endpoint was dose-dependent. A significant and dose-dependent increase of the percentage of livers showing hydropic change (Fig. 1E) was observed in NP1 (17.59 ± 7.67%) and NP2 (91.90 ± 4.74%) groups, too. Melanomacrophage centres (MMcs, Fig. 1F) have been observed only in the hepatic tissue of juveniles exposed to the highest concentration NP2 (33.47 ± 12.37%), while this parameter was not present in control and NP1. In NP exposed juveniles, degeneration of hepatic tissue was observed only in few specimens; in particular, degenerative areas were present in the livers of 6.48 ± 3.20% of NP1 and in 7.12 ± 2.57% of NP2 exposed groups. No significant differences were observed for blood vessel congestion and haemorrhages in the livers of NP exposed groups in respect to control, while ceroids were never observed. In t-OP exposed juveniles, a significant increase of the parameter loss of cord structure was observed in the hepatic tissue of both groups, with 11.23 ± 3.0% of t-OP1 and 85.71 ± 14.29% of t-OP2 livers showing hepatocytes not correctly organized in cords. The occurrence of this endpoint directly followed the concentration of t-OP. Ceroids were significantly present in the liver of t-OP2 exposed juveniles (71.43 ± 18.44%), but absent in the livers of both Cr and t-OP1 exposed groups. Haemorrhages (Fig. 1G) and

Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015

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Fig. 1. Light micrographs of Sparus aurata liver. (A) Hepatic tissue in Cr fish: hepatocytes organized in well-defined cords (arrows), sinusoids (arrowhead) and exocrine pancreatic tissue (PT) surrounding a blood vessel. (B) Some hepatocytes showing mild lipid vacuolation (arrows) in Cr; (C) some hepatocytes showing moderate lipid vacuolation (arrows); (D) all the tissue is affected by diffuse severe lipid vacuolization with hepatocyte showing a marked nuclear displacement. (E) Hydropic change. Hepatocytes appear cloudy, granular and enlarged due to a water influx. (F) Melanomacrophage centre (arrow) in NP2 exposed fish. (G) Large haemorrhage (rectangle) area in t-OP1 and (H) degeneration tissue area (circle) in t-OP2 exposed fish. A–E and G: Haematoxilin-Eosin; F and H: Mann Dominici.

hydropic change were significantly present in the livers of t-OP1 exposed juveniles (10.08 ± 3.65% and 57.66 ± 10.65%, respectively) and absent in the livers of both control and t-OP2 exposed juveniles. Tissue degeneration was present in the 100% of the livers of t-OP2 exposed juveniles (Fig. 1H). The foci of local hepatic tissue

necrosis were characterized by entirely destroyed hepatic cords and, in most cases, displayed no cellular structure. This morphological alteration was absent in the livers of Cr and only very weakly present in t-OP1 exposed (1.25 ± 1.25%) juveniles. No significant differences were observed between control and exposed juvenile

Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015

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I. Traversi et al. / General and Comparative Endocrinology xxx (2014) xxx–xxx

for Cr and NP1 groups (Fig. 3A). In t-OP exposed juveniles, the mean score for liver histological alterations showed a significant dose-dependent increase for t-OP1 (3.06 ± 0.09) and t-OP2 exposed juveniles (5.0 ± 0.0) (Fig. 3B).

Table 2 Occurrence of lipid accumulation in S. aurata juvenile liver. Absence and occurrence of lipid accumulation types (mild, moderate and severe) in control (Cr) and NP1, NP2, tOP1 and t-OP2 treated fish (n = 15 for each treatment). Values are means ± SD. Different letters indicate significant differences between groups (p < 0.05) (Bonferroni test). Groups

Cr NP1 NP2 Cr t-OP1 t-OP2

Lipid accumulation in the hepatocytes

3.3. Biochemical biomarkers – enzymatic activities

Absent

Mild

Moderate

Severe

5.56 ± 13.61a 19.45 ± 37.94a,b 56.94 ± 45.57b 5.56 ± 13.61a 0.0a 0.0a

4.63 ± 7.38a 0.0a 0.0a 4.63 ± 7.38a 0.0a 0.0a

68.52 ± 20.69a 42.59 ± 30a 14.72 ± 17.57b 68.52 ± 20.69a 41.33 ± 26.9a 42.86 ± 34.5a

21.3 ± 20.61a 37.96 ± 30.60a 33.89 ± 40.1a 21.3 ± 20.61a 58.67 ± 26.9b 57.14 ± 34.5b

Catalase activity was significantly induced in NP2 (121.98 ± 8.68 U/mg prot) exposed group in respect to the Cr (53.63 ± 6.78 U/mg prot) and NP1 (74.59 ± 5.02 U/mg prot) groups. (Fig. 4A). No significant differences were observed between Cr, t-OP1 and t-OP2 exposed groups (Fig. 4B). EROD activity was significantly inhibited in NP2 (4.03 ± 0.27 lU/mg prot) in respect to Cr (5,86 ± 0.36 lU/mg prot) and NP1 (4.38 ± 0.38 lU/mg prot) groups. No significant differences were observed between Cr and NP1 exposed group (Fig. 4C). EROD activity was significantly inhibited in t-OP1 and t-OP2 exposed groups. In particular, the lowest EROD activity was measured in t-OP1 (3.16 ± 0.15 lU/mg prot), while t-OP2 EROD activity (4.23 ± 0.27 lU/mg prot) was significantly higher than in t-OP1 and lower than Cr (Fig. 4D).

livers for MMcs, while blood vessel congestion was always absent. Presence/absence of morphological endpoints in control and exposed groups are summarized in Fig. 2A–B. The mean scores for liver histological alterations, as determined from the grading system, were 1.85 ± 0.13, 2.56 ± 0.37 and 3.46 ± 0.09 for Cr, NP1 and NP2 exposed groups, respectively. The mean score for NP2 exposed group was significantly higher than

Cr

(A)

NP1

NP2

120 c

110 100

% average presence

90 80

b

c 70 60 50

b

40

b

30 a

20 10

a a

a a

a a

a

cord structure lost

ceroids

haemorrhages

blood vessel congestion

Cr

(B)

130

c

t-OP 1

a

a

0

hydropic change

a

a

a

MMc

tissue degeneration

t-OP 2

b

120 110

b

% average presence

100

b

90 80 70 60 50 40 30

10

b

b

20 a

0 cord structure lost

a

a ceroids

a

a a

a

haemorrhages

blood vessel congestion

a

hydropic change

a

a MMc

a

a

tissue degeneration

Fig. 2. Percentage occurrence of morphological endpoints in the hepatic tissue after diet exposure to NP (A) and t-OP (B) in S. aurata males. Values are expressed as means ± SD (n = 15 for each treatment). Different letters indicate significant differences between groups (p < 0.05).

Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015

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(A)

7

5.0 4.5 4.0

a

b

N P1

NP 2

grade

3.5 3.0 2.5

a

2.0 1.5 1.0 0.5 0.0 Cr

(B)

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grade

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t-O P 1

t-OP2

Fig. 3. Evaluation of the hepatic tissues of S. aurata males exposed via diet to graded concentrations of NP (A) and t-OP (B). Values are expressed as means ± SD (n = 15 for each treatment). Different letters indicate significant differences between groups (p < 0.05).

Both NP and t-OP exposures were ineffective on the GST activity (Fig. 4E and F, respectively). The linear correlation tested between NP exposure concentrations and enzymatic activity measurement was significant (p = 0.007) and positive (r = 0.41; r2 = 0.17) for Cat and significant (p = 0.05) and negative (r = 0.30; r2 = 0.09) for EROD. No significant correlation between t-OP exposure concentrations and enzymatic activities was measured.

3.4. Molecular biomarker-Western blot Western blot analysis revealed a different VG pattern induced by the different NP and t-OP doses. In control fish any VG was detected, while in NP treated plasma, the antibody cross-reacted with five different forms of the apparent molecular weight of 110, 82, 75, 60 and 45 kDa. In particular the NP2 dose caused the highest induction of all VGs isoforms with an increase of the last two forms (60 and 45 kDa) with respect to NP1 dose. In t-OP treated fish the antibody cross-reacted only with four forms of apparent molecular weight of 110, 82, 75 and 60, the 45 kDa form was not detected. In addition t-OP2 dose induced an increase of the smaller form (60 kDa) with respect to t-OP1 dose (Fig. 5). Moreover the ZRP antibody cross reacted with two different forms of the apparent molecular weight of 60 and 55 kDa while in control fish any ZRP was detected. Also in this case the NP2 treatment caused the highest induction of ZRP isoforms (Fig. 6).

3.6. Plasma sex steroids Controls showed mean plasma E2 and T values of 1096.9 ± 229.3 pg/ml and 1589.38 ± 186.8 pg/ml, respectively. Exposure to NP induced a significant increase of E2 plasma levels only at the second concentration (2453.0 ± 246.0 pg/ml), while T plasma levels were unaffected (Fig. 7A and C). Estrogen/androgen ratio was significantly increased in NP2 exposed juveniles in respect to Cr and NP1 groups (Fig. 7E). Exposure to t-OP did not induce any alteration on E2 plasma levels, while T plasma levels were significantly decreased in both t-OP1 (117.7 ± 15.1 pg/ml) and t-OP2 (128.0 ± 21.92 pg/ml) groups (Fig. 7B and D). As a consequence, E2/T ratio was significantly increased (Fig. 7F). 3.7. Statistics Significant differences were observed between Cr and NP2 (p = 0.03) and between NP1 and NP2 (p = 0.007) groups by testing the variables all together by PERMANOVA test. SIMPER test showed the following increasing average dissimilarity order: Cr vs NP1 (31.82), NP1 vs NP2 (36.34) and Cr vs NP2 (49.60). Significant differences were observed between Cr and tOP1 (p = 0.006), between t-OP1 and t-OP2 (p = 0.007) and between Cr and t-OP2 (p = 0.009), by testing the variables all together by PERMANOVA test. SIMPER test showed the following increasing average dissimilarity order: Cr vs t-OP1 (41.36), t-OP1 vs t-OP2 (45.31) and Cr vs t-OP2 (54.73). In Table 4, variables contributing to about 75% of dissimilarity for exposure pair, are shown.

3.5. Molecular biomarker-Real Time PCR 4. Discussion The quantitative measurements of HSP70 and CATD mRNA are shown in Table 3. Livers isolated from NP1 and t-OP2 treated juveniles showed a significantly higher CATD mRNA levels with respect to control ones. In addition, the levels of HSP70 mRNA were found significantly higher only on livers from t-OP2 treated males.

In recent years, a variety of synthetic compounds with either estrogenic, androgenic or anti-androgenic activity developed: they are collectively named endocrine disrupting chemicals (EDCs) and have the ability to generate adverse health outcomes by impairing

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t- OP1

t -O P2

Fig. 4. Effects of alkylphenols on liver enzyme activities. On the left, Cat (A), EROD (C) and GST (E) activities in control (Cr), NP1 and NP2 treated fish. On the right, Cat (B), EROD (D) and GST (F) activities in control (Cr), t-OP1 and t-OP2 treated fish. Values are expressed as means ± SD (n = 15 for each treatment). Different letters indicate significant differences between groups (p < 0.05).

master physiological processes (Richter et al., 2007). Usually, EDCs have developmental effects on brain and behaviour, on male and female reproductive axis, on enzyme activity, growth, metabolism, and immune response, since they interfere with the development and/or functions of different hormonal axes (Richter et al., 2007). In the last years, the incidence of obesity and associated metabolic syndrome has been dramatically increased and several recent reports indicate that EDCs have the potential to stimulate lipid accumulation in target cells, such as adipocytes and hepatocytes (Newbold et al., 2008). In this study, lipid accumulation was always present both in the control and in exposed fish to graded concentrations (5, 50 and 100 mg/kg bw) of NP and t-OP, although with different distribution and size of lipid vesicles. Usually, large lipid accumulation in fish hepatocytes has been linked to the use of high-energy diets in aquaculture as observed in the seabream (Caballero et al., 2004). When dietary lipid or energy exceed the capacity of the hepatic cells to oxidize fatty acids, or when protein synthesis is impaired, the result is the large synthesis and deposition of triglycerides in vacuoles, leading to steatosis. More recently several authors demonstrated that lipid accumulation may occur as a result of exposure to xenobiotics

and that this morphological endpoint may be indicative of shortterm exposure to pollutants such as the herbicide clomazone and polychlorinated biphenyls, (PCBs) (Calò et al., 2010; Crestani et al., 2007; Richardson et al., 2010; Uguz et al., 2003). Lipid accumulation in the hepatocytes has been also described in Gobiocypris rarus males exposed to 3 and 10 lg/L NP (Zha et al., 2007), Fatty infiltration together with hydropic degeneration has been described in Oreochromus spilurs larvae and mother exposed to different concentrations of NP (15 and 30 lg/L) (Bin-Dohaish, 2012). This has been confirmed in our study for t-OP, since a sharp increase of severe lipid accumulation up to about 60% of the hepatic tissue was observed in the fish exposed to this chemical. A different feature has been observed in the liver of sea bream males exposed to NP where a strong lipid mobilisation was evident in about 60% of NP2 (50 mg/kg bw). It would be interesting to verify if the inhibition of lipid deposition occurs through the down-regulated expression of PPARc and LXRa (Cruz-Garcia et al., 2009) rather than the enhancement of lipid mobilisation. Hydropic change and loss of cord structure proved to be significant dose-dependent endpoints in the NP liver exposed fish. The former was the most relevant parameter well correlated with NP

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Fig. 5. Effects of alkylphenols on plasma VG levels. (A) VG proteins plasma levels in control (Cr) and NP1, NP2, t-OP1 and t-OP2 treated fish (n = 5 for each treatment). Histograms indicate mean of isoforms of different molecular weights expressed on arbitrary units. (B) VG proteins representative Western blot in control (Cr) and NP1, NP2, t-OP1 and t-OP2 treated fish.

concentration (r2 = 0.66), especially in NP2 (50 mg/kg bw) livers characterising 92% of the hepatic tissue. Hydropic change represents an adaptive cell answer to stress which can be followed by the recover of cellular activities or cell death, in the case of persisting stress conditions (Hinton and Lauren, 1990; van Dyk et al., 2007). Our data show that hydropic change is not only a timedependent but also a dose-dependent endpoint. Fish hepatocytes are arranged in double-layered cords, which in a 3-dimensional perspective, form hepatic plates, similar to those of mammalian livers. Loss of the typical hepatocyte organisation has been observed in fish exposed to hepatotoxic pollutants (Athikesavan et al., 2006; Tripathi et al., 2011). In our study, loss of cord structure has been showed in a dose-dependent manner for both NP and t-OP exposed fish. Loss of normal liver tissue organization has been described in rainbow trout after 3 and 4 week exposure to 220 lg NP/L (Uguz et al., 2003), in Gobiocypris rarus after 4 week exposure to 30 lg NP/L (Zha et al., 2007), in Orechromus spilurs larvae and mothers exposed to 15 and 30 lg NP/L (Bin-Dohaish, 2012) and in C. dimerus exposed to a nominal concentration of 150 lg t-OP/L for 4 weeks (Genovese et al., 2012). The increasing dose of t-OP in the diet played a significant role on the occurrence of the histological alterations. In particular, t-OP1 (5 mg/kg bw) exposure induced haemorrhages and hydropic change, while t-OP2 (50 mg/kg bw) exposure induced the presence of ceroids and significantly increased the degeneration of the hepatic tissue. Ceroids have been usually described as a lipoid liver disease and considered as indicator of vitamin deficit (Mumford et al., 2007). The widespread tissue degeneration, observed in 100% of the t-OP2 (50 mg/kg bw) exposed fish, could be the reason of the

Fig. 6. Effects of alkylphenols on plasma ZRP levels. (A) ZRP proteins plasma levels in control (Cr) and NP1, NP2, t-OP1 and t-OP2 treated fish (n = 5 for each treatment). Histograms indicate mean of isoforms of different molecular weights expressed on arbitrary units. (B) ZRP proteins representative Western blot in control (Cr) and NP1, NP2, t-OP1 and t-OP2 treated fish.

Table 3 Effects of alkylphenols on CATD and HSP70 gene expression. Control (Cr) and NP1, NP2, t-OP1 and t-OP2 exposed fish (n = 5 for each treatment) gene expressions are presented as mean ACT and EF1a normalized copy numbers ± SD. Means in the same row with different letters are significantly different (p < 0.05) (Bonferroni test). Target gene

Cr

NP1

NP2

t-OP1

t-OP 2

CATD HSP70

2.13 ± 0.93a 1.79 ± 0.37a

8.43 ± 1.42b 1.74 ± 0.29a

3.27 ± 1.18a 1.86 ± 0.28a

5.09 ± 2.39a 1.59 ± 0.92a

7.80 ± 1.36b 4.25 ± 0.91b

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Cr

Fig. 7. Effects of alkylphenols on sex steroids. On the left, E2 (A), T (C) plasma levels and E2/T ratio (E) in control (Cr), NP1 and NP2 treated fish; on the right E2 (B), T (D) plasma levels and E2/T ratio (F) in control (Cr), t-OP1 and t-OP2 treated fish. Values are expressed as mean ± SD, n = 15 for each treatment. Different letters indicate significant differences between groups (p < 0.05).

Table 4 Dissimilarity percentage in S. aurata exposed groups. List of the principal variables contributing to the percentage dissimilarity between couple of groups. Data are expressed as percentage (SIMPER test). Contribution%

Cumulative%

Contribution%

Cumulative%

Cr vs NP1 average dissimilarity = 31.82 Moderate Cat Cord structure loss Hydropic change

21.8 18.65 18.23 15.22

21.8 40.45 58.68 73.9

Cr vs t-OP1 average dissimilarity = 41.36 Hydropic change Severe Cat Cord structure lost

28.29 19.12 11.75 11.12

28.29 47.41 59.16 70.28

Cr vs NP2 average dissimilarity = 49.60 Hydropic change Moderate Cord structure loss MMc

28.57 18.95 17.81 12.53

28.57 47.52 65.33 77.86

Cr vs t-OP2 average dissimilarity = 54.73 Tissue degeneration Cord structure lost Ceroids Severe

24.27 20.47 16.39 13.89

24.27 44.74 61.12 75.02

NP1 vs NP2 average dissimilarity = 36.34 Hydropic change Moderate MMc Cord structure loss

26.9 17.53 16.39 14.97

26.9 44.44 60.83 75.79

t-OP1 vs t-OP2 average dissimilarity = 45.31 Tissue degeneration Hydropic change Ceroids Cord structure lost

23.2 16.43 16.38 15.81

23.2 39.63 56.01 71.82

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lack of endpoints such as haemorrhages, hydropic change, and MMcs in the livers of this group. The liver of fish is responsible for the same basic metabolic functions as in mammals, included the excretion and metabolism of xenobiotic compounds. Fish have analogous mechanisms for handling xenobiotic compounds, including both phase1 and phase2 biotransformation reactions (Cowey and Walton, 1989). It is well known that the exposure to NP can inhibit liver EROD activity in fish (Teles et al., 2005; Vaccaro et al., 2005; Pérez-Carrera et al., 2007). In agreement with these studies, our results show a significant inhibition of this enzymatic activity, starting from the second concentration (NP2). This is also confirmed by the linear correlation (significant and negative) between exposure concentration and EROD activity. We found the same significant inhibitory effect observed for the NP, in t-OP exposed fish too, but with an U-shaped profile. This is in contrast with the results obtained in the rainbow trout hepatocytes exposed to t-OP where EROD activity was increased, even though not significantly (Navas and Segner, 2000). Several studies demonstrated the induction of Glutathione-S-Transferase (GST) activity after exposure to both NP and t-OP not only in S. aurata (Pérez-Carrera et al., 2007; Teles et al., 2005), but also in other species such as Siganus oramin (Du et al., 2008). In this study we did not found any alteration in the GST activity after exposure to both alkylphenols, in agreement with Sturvè et al. (2006) in Gadus morhua. Catalase activity (Cat) is a very known biomarker of oxidative stress (Borkovic´ et al., 2005). In the study of Schlezinger et al. (2006) chemicals can inhibit EROD activity, through uncoupling of cytochrome P450 1A with redox oxygen species (ROS) release that could exert induction on enzymes like Cat. The ROS release could explain our results after exposure to NP where a significant induction of Cat activity was observed in concomitance with an EROD inhibition. The stressor effect of NP and t-OP at hepatic level is supported by the dose related increase of HSP70 gene expression, confirming the higher t-OP toxicity. These contaminants also affects the VG and ZRP synthesis in male concomitant with the increase of CATD, a key enzyme in the VG processing bringing to yolk protein formation. Since, both VG and CATD genes expression are regulated by and oestrogenresponsive element (ERE) in the promoter (Teo et al., 1998), these results confirm the estrogenic effects of these two contaminants, and the risk associated with the impairment of reproduction when these contaminants are present in the food chain with risk for consumers that are under constant exposure to these contaminants. To supplement the morphological endpoints and molecular biomarkers, this study has focused on sex steroid plasma levels, considering that a balance between estrogens and androgens is necessary to the correct organ development and functioning. It is known that numerous EDCs can exert their effects through the modulation of steroid pathways (Jalabert et al., 2000). The increase of estrogens together with a decrease of androgens has been observed in Carassius auratus juveniles exposed to NP and explained by the induction of the aromatase activity (Soverchia et al., 2005). In our study, the dose-dependent increase of E2 plasma levels did not match with a significant corresponding decrease of T, as also observed in immature male yellowfin seabream (Naderi et al., 2012). In t-OP exposed S. aurata males, T plasma level has been significantly decreased as observed in juvenile male summer flounder exposed for 4 weeks to 200 mg:kg t-OP (Mills et al., 2001). The same has been described in Siganus oramin juveniles fed with increasing concentrations of t-OP for 28 days by Du et al. (2008) that considered this effect as a consequence of deregulation of the hypothalamo–pituitary–gonad axis or blocking of hydroxylase activity (Hanioka et al., 2000). Our results show that NP and t-OP can exert estrogenic and anti-androgenic effects on S. aurata males, respectively. Another way to evaluate sex steroid

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hormones is to examine their ratio. Information on normal levels of sex steroid ratio is scarce and ranges have not been established for classifying fish as normal (healthy) or not, nevertheless the ratio of oestrogen to androgen has been suggested as a good marker of abnormal sex steroid concentrations (Goodbred et al., 1997). In this case, steroid ratio was not able to discriminate between NP and t-OP exposed groups. In conclusion, histological analysis confirmed to be a very sensitive and crucial parameter in determining cellular changes generated by pollutant exposure that may occur in target organs, such as the liver (Hinton et al., 2001). The histological analysis applied on S. aurata liver provided a quantitative assessment of the lesions and, with lowest dose, showed a greater toxic effect of t-OP respect to NP. Within the battery of endpoints considered in this study, Cat was the only enzymatic activity showed by SIMPER analysis amongst the principal variables of dissimilarity. The different effects exerted by the exposure to the two contaminants are clearly highlighted by the score obtained by the endpoints using morphological approach. Acknowledgments This work was funded by the Italian Ministry of Health, project ‘‘Food and environmental safety: the problem of the endocrine disruptors’’ contract Nr. RF-2009-153685. The authors thank very much Prof. Maria Chiara Chiantore and Dr. Irene Schiavetti for their help in the statistical processing of data. References Ackermann, G.E., Swaiger, J., Negele, R.D., Fent, K., 2002. Effects of longterm nonylphenol exposure on gonadal development and biomarkers of estrogenicity in juvenile rainbow trout (Oncorhynchus mykiss). Aquat. Toxicol. 60, 203–221. Athikesavan, S., Vincent, S., Ambrose, T., Velmurugan, B., 2006. Nickel induced histopathological changes in the different tissues of freshwater fish, Hypophthalmichthys molitrix (Valenciennes). J. Environ. Biol. 27, 391–395. Augereau, P., Miralles, F., Cavaillès, V., Gaudelet, C., Parker, M., Rochefort, H., 1994. Characterization of the proximal estrogen-responsive element of human cathepsin D gene. Mol. Endocrinol. 8, 693–703. Barret, A.J., 1977. Cathepsin D and other carboxyl proteinases. In: Barret, A.J. (Ed.), Proteinases in Mammalian Cells and Tissues. Biomedical Press, Elsevier, London, pp. 209–248. Bergmeyer, H.U., Grassl, M., Walter, H.E., 1983. Methods of enzymatic analysis. In: Bergmeyer, H.U. (Ed.), VCH Weinheim, vol. 2. Deerfield Beach, FL, p. 249. Bin-Dohaish, E.A., 2012. The effects of 4-nonylphenol contamination on livers of Tilapia fish (Oreochromus spilurs) in Jeddah. Biol. Res. 45, 15–20. Bjerregaard, P., Andersen, S.B., Pedersen, K.L., Pedersen, S.N., Korsgaard, B., 2007. Orally administered bisphenol a in rainbow trout (Oncorhynchus mykiss): estrogenicity, metabolism, and retention. Environ. Toxicol. Chem. 26, 1910– 1915. Borkovic´, S.S., Šaponjic´, J.S., Pavlovic´, S.Z., Blagojevic´, D.P., Miloševic´, S.M., Kovacˇevic´, T.B., Radojicˇic´, R.M., Spasic´, M.B., Zˇikic´, R.V., Saicˇic´, Z.S., 2005. The activity of antioxidant defence enzymes in the mussel Mytilus galloprovincialis from the Adriatic Sea. Comp. Biochem. Physiol. 141C, 366–374. Booth, J., Boyland, E., Sims, P., 1961. An enzyme from rat liver catalysing conjugations with glutathione. Biochem. J. 79, 516–524. Bouzas, A., Aguado, D., Martí, N., Pastor, J.M., Herráez, R., Campins, P., Seco, A., 2011. Alkylphenols and polycyclic aromatic hydrocarbons in eastern Mediterranean Spanish coastal marine bivalves. Environ. Monit. Assess. 176, 169–181. Bradford, M.M., 1976. A rapid and sensitive assay of protein utilizing the principle of dye binding. Anal. Biochem. 72, 248–254. Caballero, M.J., Izquierdo, M.S., Kjørsvik, E., Fernàndez, A.J., Rosenlund, G., 2004. Histological alterations in the liver of sea bream, Sparus aurata L., caused by short- or long-term feeding with vegetable oils. Recovery of normal morphology after feeding fish oil as the sole lipid source. J. Fish Dis. 27, 531– 541. Calò, M., Albergina, D., Bitto, A., Lauriano, E.R., Lo, Cascio P., 2010. Estrogenic followed by anti-estrogenic effects of PCBs exposure in juvenil fish (Sparus aurata). Food Chem. Toxicol. 48, 2458–2463. Carnevali, O., Maradonna, F., 2003. Exposure to xenobiotic compounds: looking for new biomarkers. Gen. Comp. Endocrinol. 131, 203–209. Carrera, E.P., García-López, A., Martín del Río, P., Martínez-Rodríguez, G., Solé, M., Mancera, J.M., 2007. Effects of 17beta-estradiol and 4-nonylphenol on osmoregulation and hepatic enzymes in gilthead sea bream (Sparus auratus). Comp. Biochem. Physiol. C: Toxicol. Pharmacol. 145, 210–217.

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Please cite this article in press as: Traversi, I., et al. Alkylphenolic contaminants in the diet: Sparus aurata juveniles hepatic response. Gen. Comp. Endocrinol. (2014), http://dx.doi.org/10.1016/j.ygcen.2014.06.015