Responses of conventional and molecular biomarkers in turbot Scophthalmus maximus exposed to heavy fuel oil no. 6 and styrene

Aquatic Toxicology 116–117 (2012) 116–128

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Responses of conventional and molecular biomarkers in turbot Scophthalmus maximus exposed to heavy fuel oil no. 6 and styrene Pamela Ruiz a , Maren Ortiz-Zarragoitia a , Amaia Orbea a , Michael Theron b , Stéphane Le Floch c , Miren P. Cajaraville a,∗ a

Laboratory of Cell Biology and Histology, Faculty of Science and Technology, University of the Basque Country, Sarriena z/g, E- 48940 Leioa, Basque Country, Spain Laboratoire ORPHY, EA 4324, UFR Sciences et Techniques, Université de Bretagne Occidentale, 6 avenue Le Gorgeu, CS 93837, 29238 Brest Cedex 3, France c Centre of Documentation, Research and Experimentations on Accidental Water Pollution, 715 rue Alain Colas, CS 41836, 29218 Brest Cedex 2, France b

a r t i c l e

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Article history: Received 3 October 2011 Received in revised form 17 January 2012 Accepted 5 February 2012 Keywords: Conventional and molecular biomarkers Heavy fuel oil no. 6 Styrene Recovery Turbots

a b s t r a c t Several accidental spills in European coastal areas have resulted in the release of different toxic compounds into the marine environment, such as heavy fuel oil type no. 6 in the “Erika” and “Prestige” oil spills and the highly toxic styrene after the loss of the “Ievoli Sun”. There is a clear need to develop tools that might allow assessing the biological impact of these accidental spills on aquatic organisms. The aim of the present study was to determine the short-term effects and recovery after exposure of juvenile fish (Scophthalmus maximus) to heavy fuel oil no. 6 and styrene by using a battery of molecular, cell and tissue level biomarkers. Turbots were exposed to styrene for 7 days and to the diluted soluble fraction of the oil (10%) for 14 days, and then allowed to recover in clean seawater for the same time periods. cyp1a1 transcript was overexpressed in turbots after 3 and 14 days of exposure to heavy fuel oil, whereas ahr transcription was not modulated after heavy fuel oil and styrene exposure. ppar˛ transcription level was significantly up-regulated after 3 days of treatment with styrene. Liver activity of peroxisomal acylCoA oxidase (AOX) was significantly induced after 14 days of oil exposure, but it was not affected by styrene. Hepatocyte lysosomal membrane stability (LMS) was significantly reduced after exposure to both treatments, indicating that the tested compounds significantly impaired fish health. Both AOX and LMS values returned to control levels after the recovery period. No differences in gamete development were observed between fuel- or styrene- exposed fish and control fish, and vitellogenin plasma levels were low, suggesting no xenoestrogenic effects of fuel oil or styrene. While styrene did not cause any increase in the prevalence of liver histopathological alterations, prevalence of extensive cell vacuolization increased after exposure to heavy fuel oil for 14 days. In conclusion, the suite of selected biomarkers proved to be useful to determine the early impact of and recovery from exposure to tested compounds in turbot. © 2012 Published by Elsevier B.V.

1. Introduction Accidental spills in the marine environment cause significant chemical pollution, with acute effects on exposed organisms (Kirby and Law, 2010). A large number of marine ecosystems have been affected by spills. The “Exxon Valdez” oil spill caused the release of 42,000 tons of crude oil in Alaska in 1989 (Spies et al., 1996); the

Abbreviations: AHR, aryl hydrocarbon receptor; AOX, peroxisomal acyl-CoA oxidase; CYP1A1, cytochrome P4501A1; EF1-␣, elongation factor 1 alpha; EPA, U.S. Environmental Protection Agency; EROD, ethoxyresorufin O-deethylase; LP, labilization period; MMC, melanomacrophage center; PAH, polycyclic aromatic hydrocarbon; PPAR, peroxisome proliferator-activated receptor; RQ, relative quantification; VTG, vitellogenin; WSF, water soluble fraction. ∗ Corresponding author. Tel.: +34 94 6012697; fax: +34 94 6013500. E-mail address: [email protected] (M.P. Cajaraville). 0166-445X/$ – see front matter © 2012 Published by Elsevier B.V. doi:10.1016/j.aquatox.2012.02.004

tanker “Erika” lost 20,000 tons of fuel oil (classified as no. 2 according to Association Franc¸aise de Normalisation and no. 6 according to the American Society for Testing and Materials) in front of the coast of Brittany in 1999 (Bocquené et al., 2004); the sinking of the tanker “Ievoli Sun” resulted in the release of more than 1000 tons of the hydrocarbon styrene in the English Channel in 2000 (Law et al., 2003); the tanker “Prestige” sunk in front of the Galician coast in 2002 spilling more than 60,000 tons of heavy fuel oil (González et al., 2006), and more recently the spill which occurred in the Gulf of Mexico released 780,000 tons of oil to the sea impacting the Southeast coastline of USA (Mitsch, 2010). The spilled compounds may affect the organisms producing changes at molecular, cellular and physiological levels which at long-term may provoke the decline of their populations, as shown after the “Exxon Valdez” oil spill (Peterson, 2001). Several works have employed biomarkers to assess biological effects after

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accidental spills. Ten years after the “Exxon Valdez” oil spill, exposure of fish to oil was still detected by means of ethoxyresorufin O-deethylase (EROD) activity and biliary fluorescent metabolites (Huggett et al., 2003; Jewett et al., 2002). Likewise, 1 year after the “Prestige” oil spill peroxisomal acyl-CoA oxidase (AOX) activity, lysosomal responses such as changes in the structure and membrane stability and histopathology in mussel and fish discriminated the most impacted areas by the spilled oil (Marigómez et al., 2006; Orbea et al., 2006). Laboratory studies have also been carried out to assess the biological effects of polycyclic aromatic hydrocarbons (PAHs) in aquatic animals. PAHs are the main components of these spilled oils (Alzaga et al., 2004) and are able to induce biotransformation metabolism and peroxisome proliferation in marine organisms (Cajaraville et al., 2003; Bilbao et al., 2010). In vertebrates, biotransformation and peroxisome proliferation are mediated by aryl hydrocarbon receptor (AhR) and by peroxisome proliferatoractivated receptors (PPARs) (Barron et al., 2004; Mandard et al., 2004), respectively. PAHs are also the oil components that pose the highest environmental risk, mainly due to their carcinogenic and mutagenic properties (Beyer et al., 2010). PAHs are absorbed by fish via the gills and body surface, but also by ingestion of food or through contaminated sediment (Van der Oost et al., 2003). Compared to other aromatic compounds, little attention has been paid to styrene, probably due to its lower rate of industrial use in comparison with PAHs (Cushman et al., 1997; Gibbs and Mulligan, 1997; Mamaca et al., 2005). Styrene is primarily used in the production of polymers and polystyrene. Styrene is commonly shipped by vessel and due to its characteristics (volatility and low solubility in water) it has rarely been detected in water and, when present, it occurs at very low levels (ng L−1 ) (Cushman et al., 1997). Based on the results of toxicological studies and on measured concentrations in water, styrene is not deemed to cause effects on aquatic organisms as a consequence of environmental exposure, except in the immediate vicinity of a spill (Fu and Alexander, 1992; Cushman et al., 1997). The main objective of the present study was to determine the short-term effects of and recovery from laboratory exposure of juvenile turbots (Scophthalmus maximus) to two compounds appearing in the environment as a result of accidental spills, the complex mixture formed by heavy fuel oil no. 6, rich in PAHs, and the highly volatile hydrocarbon, styrene. The experimental system was designed to simulate a realistic scenario immediately after a spill. Juvenile turbots were exposed to the water soluble fraction of fuel oil no. 6 for 14 days and to styrene for 7 days, and then transferred to clean seawater for the same period of time. A longer exposure period for styrene would not be realistic or relevant, because styrene evaporates quickly in open systems and, thus, high concentrations of this compound would not last in the aquatic environment (Alexander, 1997). On the other hand, oil components are detected in the water column for longer periods after a spill and, thus, exposure to oil was extended for 2 weeks. During the exposure periods, PAH and styrene levels were monitored in seawater and, at the end of the experiment, styrene concentration in tissues was quantified to evaluate bioaccumulation processes. An integrated battery of biomarkers from molecular to cell and tissue levels were measured which reflects impact at different levels of biological organization. Peroxisome proliferation and induction of biotransformation metabolism were studied as exposure biomarkers for organic pollutants. Peroxisome proliferation was assessed at gene transcription level (ppar˛) and at AOX activity level (Cajaraville et al., 2003; Zorita et al., 2008). Induction of biotransformation metabolism was studied by ahr and cyp1a1 transcription level (Hahn, 2002; Zhou et al., 2005). Lysosomal membrane destabilization in fish hepatocytes was used as an indicator of non-specific effects arising from an increased intracellular accumulation of xenobiotics (Köhler, 1991). Plasma

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vitellogenin levels were measured as biomarker of estrogenicity (Arukwe and Goksøyr, 2003) and gonad histology was also evaluated as supporting parameter. Finally, livers were examinated for histopathological alterations since several studies carried out in coastal waters have shown correlation between environmental contaminants and the occurrence of toxicopathic liver lesions in fish (Feist et al., 2004; Stentiford et al., 2003; Vethaak et al., 1996). In recent years, fish diseases and liver histopathological alterations have been used as indicators of pollution effects and have been implemented in monitoring programs (Feist et al., 2004; Lang, 2002). The presence of inflammatory lesions, hepatocellular fibrillar inclusions, and preneoplastic and neoplastic lesions is higher in fish captured in polluted environments than in fish from reference sites (Stentiford et al., 2003). 2. Materials and methods 2.1. Animals Juvenile turbots, S. maximus (n = 500; weight 358 ± 93 g, length 28 ± 2 cm, mean ± SD) were purchased from a fish farm (France Turbot, Tredarzec, France). Before starting the experiment, they were acclimatized for 20 days in two laboratory tanks of 1000 L (control tank and treatment tank). They were fed daily with dried pellets (Aquaculture food Le Gouessant® , Lamballe, France, 4.5 mm diameter, total protein 54% of dry matter and crude fat 12% of dry matter). Running seawater was supplied at a flow rate of 5 L min−1 in both tanks. Water was taken from Sea of Iroise, France, and sterilized by means of UV-rays and filters system. Experiments were performed in the facilities of Cedre (Brest, France). 2.2. Experimental systems 2.2.1. Exposure of turbots to fuel oil number 6 Experiments were carried out in November 2006. The light regime was set according to the season: 14 h light, 10 h dark. Water salinity (35–36‰), water pH (7.98 ± 0.06), oxygen concentration (228 ± 18 ␮mol L−1 ) and sea water temperature (16 ± 2 ◦ C) were measured daily. During the experiment, turbots were fed twice per day with dried pellets. Turbots were exposed for 14 days to 10% of the water soluble fraction (WSF) of heavy fuel oil no. 6 using a continuous flowthrough system with water aeration described by Aas et al. (2000). The system made it possible to expose fish to relative stable concentrations of dispersed crude oil (Aas et al., 2000), similar to those reported after accidental oil spills such as Prestige and Exxon Valdez shipwrecks: 2090 ng L−1 (González et al., 2006) and 6000 ng L−1 (Short and Harris, 1996), respectively. The experimental system consisted of a 2 m high column, which was filled with glass beads coated with oil and seawater was running through the column from top to bottom, to produce the WSF of the oil, which was diluted ten times before being administered to the animals. After the exposure period, fish were maintained for 14 additional days in clean seawater for recovery. A sample of heavy fuel oil was collected to characterize the PAH composition of fuel oil used in the experiment. Seawater samples were taken at T0 (before fish were introduced), and after 1, 2, 3, 7 and 14 days of exposure to monitor PAH concentration in the water column of the contamination tank along the exposure period. Fish samples were collected before exposure started (T0) and at 3, 7 and 14 days of exposure. A last sampling was done after additional 14 days of recovery in clean seawater. At each sampling time of the exposure period, 12 fish per experimental group were sacrificed. From these samples, a piece of liver of 7–8 turbots per experimental group was immersed individually

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Fig. 1. Schematic representation of the styrene exposure system.

in RNA later® (Sigma–Aldrich, St. Louis, Missouri, USA) and frozen in liquid nitrogen for transcription studies. Livers, gonads and blood samples of 12 turbots per experimental group were collected and processed individually at each sampling time for AOX activity, plasma vitellogenin level, and gonad and liver histopathological analysis. For lysosomal membrane stability test, a small piece of the liver obtained from 5 fish per experimental group were rapidly frozen in liquid nitrogen and stored at −80 ◦ C until sectioning. After the recovery period, 5 fish previously exposed to fuel oil and 12 control fish were sacrificed and samples for the all above determinations were collected from each individual. 2.2.2. Exposure of turbots to styrene Turbots were exposed to styrene for 7 days. Then fish were maintained for another 7 days in clean water for the recovery. The styrene exposure system was built in order to perform a controlled exposure to the chemical in a realistic way, simulating the possible concentration present in the sea surface after a styrene spill. The experimental system was composed of a mixing tank connected to an exposure tank and a degassing column (Fig. 1). In the mixing tank, which had a capacity of 316 L, a slick of styrene was placed in contact with the aqueous phase and dissolved slowly. Then, contaminated water containing the dissolved fraction of the slick was pumped to the exposure tank (300 L) where turbots were maintained. Styrene disappeared rapidly from water and, therefore, the concentration remaining in the exposure tank would be similar to that appearing in seawater immediately after a spill. The exposure tank was connected to a system of waste water recuperation by overflow. During the experimentation, a water flow of 1.3 L min−1 was pumped and sent on a degassing column before being directed to the mixing tank. A second circuit ensured a renewal of fresh sea water to maintain the nitrate and nitrite concentrations stable. In addition, the oxygen saturation of water was maintained into this tank by a compressor, which injected air via diffuser. The degassing column favored the gaseous

exchange, allowed water oxygenation by percolation on glass beads and allowed elimination of dissolved carbon dioxide produced by organisms (carbon dioxide transfer from water to air). During the experimentation, both mixing and exposure tanks were kept covered to avoid cross contamination from the styrene exposure tank to the control tank. The concentration of styrene was monitored in seawater during the experiment. Fish samples were collected before the exposure (T0), after 3 and 7 days of exposure and after 7 days of recovery in clean water. Muscle tissues of 10 control and 10 styrene-exposed turbots were collected after 7 days of exposure to determine the level of styrene in fish. Styrene concentration was measured in muscle because xenobiotic metabolism in this organ is much slower than in liver and, thus, styrene accumulation in muscle could better reflect environmental exposure. Livers of 6-7 turbots per experimental group at each sampling time were dissected out, immersed individually in RNA later® and frozen in liquid nitrogen for transcription studies. Livers, gonads and blood samples of 12 turbots per experimental group were collected and processed individually at each sampling time for liver AOX activity, plasma vitellogenin levels and histopathological analysis of liver and gonad. For lysosomal membrane stability test a small piece of the liver obtained from 5 fish per experimental group were rapidly frozen in liquid nitrogen and stored at −80 ◦ C until sectioning. 2.3. Chemical analyses 2.3.1. PAH concentrations in heavy fuel oil no. 6 and seawater during exposure The composition of the oil used in this study was determined according to Wang and Fingas (1995). The 16 priority PAHs of the U.S. Environmental Protection Agency (EPA) were measured (naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benzo[a]anthracene, chrysene, benzo[b]fluoranthene,

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benzo[k]fluoranthene, benzo[a]pyrene, indeno[1,2,3-c,d]pyrene, dibenz[a,h]anthracene and benzo[g,h,i]perylene). Around 20–30 mg of the heavy fuel oil was spiked with 100 ␮L of a mixture of internal standards (perdeuterated PAHs) and purified by solid phase extraction (cyanopropylphase, compounds eluted by 5 mL of a mixture pentane/dichloromethane 80/20). The sample volume was then reduced to 200 ␮L prior to the gas chromatography coupled to mass spectrometry (GC–MS) analysis. The extracted compounds were analyzed on a HP 5890 series II gas chromatograph coupled to a HP 5973 mass selective detector (MSD) (Electronic impact: 70 eV, voltage: 1300 V) (Hewlett-Packard, Palo Alto, CA, USA). Samples were injected in the split/splitless injector into a 30 m × 0.25 mm × 0.25 mm HP 5-MS capillary column (Hewlett-Packard, Palo Alto, CA, USA). The GC temperature gradient was from 50 ◦ C (1 min) to 300 ◦ C (20 min) at 5 ◦ C min−1 . The carrier gas was helium and was kept at a constant flux of 1 mL min−1 . PAH quantification was conducted using selected ion monitoring mode with the molecular ion of each compound at a minimum of 1.5 scan s−1 . PAHs were quantified relative to the perdeuterated PAHs introduced at the beginning of the sample preparation procedure. Individual perdeuterated PAHs (naphthalene d8, biphenyl d10, phenanthrene d10, chrysene d12 and benzo[a]pyrene d12 at concentrations of, respectively, 210, 110, 210, 40 and 40 mg mL−1 in acetonitrile) were purchased from Sigma–Aldrich (France). Calibration curves were established using a mixture of 16 parent PAHs purchased from LGC Standards (France), each compound at the concentration of 100 mg mL−1 . For determining PAHs concentration in sea water, water samples (1 L) were collected into Duran glass bottles that had been previously heated at 500 ◦ C to eliminate impurities. Samples were extracted using Pestipur grade dichloromethane (3 × 100 mL per seawater sample). The combined organic extracts were dried by filtering through anhydrous sodium sulphate (Na2 SO4 Pestipur grade) and concentrated to 2 mL by means of a Turbo Vap 500 concentrator (Zyman, Hopkinton, MA, USA). Aromatic compounds were analyzed using the same GC–MS protocol previously described. The detection limit of this method was 1 ng L−1 . 2.3.2. Water and tissue analyses of styrene Styrene concentration in water was directly monitored by UVspectrofluorometer (SF-UV) quantified by headspace (HS)/GC–MS using the gerstel static HS module mounted on automated sampler. The combination of these two analytical techniques allowed to follow the styrene concentration in a qualitative way throughout the exposure and to evaluate the exposure concentration. Ten milliliter water samples were stirred for 10 min at 70 ◦ C. Styrene was then transferred to a HP-5MS column (Hewlett-Packard, Palo Alto, CA, USA) using a 2.5 mL syringe. The split/splitless injector was used in split mode at 250 ◦ C. The oven program of temperature was the following: from 40 ◦ C (2 min) to 100 ◦ C at 7 ◦ C min−1 , and, then to 150 ◦ C (2 min) at 25 ◦ C min−1 . The mass spectrometer was operated in SIM mode as previously. Styrene was quantified with respect to ethyl benzene using a calibration curve (from 0 to 10 mg L−1 ). For tissues analysis, 1 mL of methanolic solution of ethyl benzene was added to 3 g of muscle per each fish. Then, sample was placed in the flask of the HS module and the same GC–MS protocol was used. 2.4. Sequencing of target sequences and quantitative real-time RT-PCR For sequencing studies, juvenile turbots were collected in a farm in San Sebastian, Basque Country, Spain. Livers were frozen in RNA later® and stored at −80 ◦ C until processing. Fifty to a hundred mg of liver were homogenized in TRIzol® (Invitrogen, Carlsbad,

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California, USA) using a Hybaid RyboliserTM (Hybaid, Ashford, UK) cell disruptor at a shaking speed of 4 m s−1 for 40 s. Total RNA was isolated for each individual fish liver following the manufacturer’s instructions. Then, total RNA was used as template for cDNA synthesis by Super ScriptTM II reverse transcriptase PCR (Invitrogen, Leek, Netherlands) using random hexamers as primers and following the manufacturer’s recommendations in a conventional iCycler termocycler (Bio-Rad, California, USA). cDNAs were PCR amplified using degenerate primers, Fw-5 -CCA CAC TGA GAC CAG CAG-3 and Rv-5 -G(GCT)T TGT G(C,T)T TGG TCC TGA AGA-3 for ahr and Fw-5 GAT GGA GCC CAA GTT (AG)CA G-3 and Rv-5 - CTT GAT TTC CTG CAC (GC)AG C-3 for ppar˛, designed through ClustalW alignments of known teleost and other phyla sequences to amplify conserved regions of ahr and ppar˛. These sequences were retrieved from the National Center for Biotechnology Information (NCBI, U.S. National Library of Medicine, USA). PCR was done with a 2 min activation and denaturing step at 94 ◦ C, followed by 35 cycles of a 30 s denaturing step at 94 ◦ C and a 30 s extension step at 72 ◦ C. The annealing temperature was 57 ◦ C for ahr and 55 ◦ C for ppar˛. PCR products were visualized in a 1.5% agarose gel, stained with ethidium bromide and purified using a PCR purification kit (Qiagen, Hilden, Germany). Single bands of ppar˛ were directly sequenced, but whenever more than one band was amplified for ahr, purified products were cloned using the TOPO-TA cloning kit (Invitrogen, Carlsbad, California, USA). PCR products for each gene were sequenced using degenerate primers while cloned inserts were sequenced using universal M13 primers in the Sequencing and Genotyping Service (SGIker) of the University of the Basque Country. Alignments and similarity matrices were performed using Blastn, Blastx and ClustalW. Transcription levels of cyp1a1 (GenBank ID: AJ310694), ahr and ppar˛ were measured in turbot livers by real time quantitative PCR using TaqMan probes and primers. About 50–100 mg of fish liver was homogenized in TRIzol® . Total RNA was isolated and cDNA was obtained from 1 ␮g of total RNA by Super ScriptTM II reverse transcriptase PCR (Invitrogen, Leek, Netherlands) using random hexamers as primers and following the manufacturer’s recommendations in the iCycler termocycler. The real time PCR was run in 25 ␮L reactions on a 7003 PCR machine (Applied Biosystems, California, USA) using TaqMan Reverse Transcription Reagent (New Jersey, USA). TaqMan probes and primers from turbot specific sequences were designed using Primer Express 3.0 software (Applied Biosystems, California, USA). All details are presented in Table 1. Universal conditions were used in PCR for all genes: 1 cycle at 50 ◦ C for 2 min, 1 cycle at 95 ◦ C for 10 min, 40 cycles at 95 ◦ C for 15 s and at 60 ◦ C for 1 min. Amplified fragments were visualized after PCR in ethidium bromide stained 1.5% agarose gels. Then, bands were cloned by TOPO-TA cloning kit and sequences were confirmed by sequencing. Elongation factor 1 alpha (EF1-˛, GenBank ID: AF467776) was used for normalization of transcription levels of target genes. Relative transcript expression of a gene was calculated with the 2−ct method (Livak and Schmittgen, 2001) relative to the mean of control animals sampled at day 3. Similar results were obtained when the mean value of the control group for each sampling day was used, but this approach did not allow studying the time-related variation in the transcription levels of control animals. 2.5. Acyl-CoA oxidase (AOX) activity Livers were homogenized in 4 mL of TVBE buffer (1 mM sodium bicarbonate, 1 mM EDTA, 0.1% ethanol and 0.01% Triton X-100, pH 7.6) per gram of tissue, using a glass-Teflon homogenizer held in an ice bath. Homogenates were centrifuged at 500 g at 4 ◦ C for 15 min, and obtained supernatants were diluted 1:10 in TVBE buffer

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Table 1 5 –3 forward (Fw) primers, 5 –3 reverse (Rv) primers, and 5 –3 dual label probes (Probe) with indicated fluorophore reporter molecule (FAM) and the quencher NFQ dye used for TaqMan real time PCR of studied genes. Gene (GenBank accession no)

Fw

Rv

Probe

Product size (bp)

cyp1a1 (AJ310694) ahr (JN253594) ppar˛ (JN253593) EF-˛ (AF467776)

CTGCAAAGAGGGAGAGTATCTGAT TGGCGATGGCGAAGAGA AGAACATTGTGCAGGTGCTACAG CCCCGGCGACAACGT

GTTAGCCACAGACACAACAATGTAG TGTGTGACGACGAGGATGGT ACGTGTCGTCCGGATGGT CGCGGCGGATCTCCTT

FAM-CAGCTTCGACCCCTTCC-NFQ FAM-CGAGCTGATGTGGCG-NFQ FAM-TCCACCTGCTGGCC-NFQ FAM-AACATCAAGAACGTGTCCGT-NFQ

100 53 56 59

and assayed for AOX activity. AOX activity was determined spectrophotometrically measuring the H2 O2 -dependent oxidation of dichlorofluorescein diacetate catalyzed by an exogenous peroxidase using palmitoyl-CoA 30 ␮M as substrate (Small et al., 1985). Measurements ( = 502 nm) were performed for four min during the linear phase of the reaction. Each sample was measured at least twice and values accepted if the difference between them was less than 10%. Total protein concentration was measured using the DC Protein Assay of Bio-Rad, based on the method of Lowry et al. (1951), with gamma-globulin as standard. AOX activity is given as mU AOX mg−1 protein. 2.6. Lysosomal membrane stability At least 12 serial tissue sections (10 ␮m thick) of a small piece of the liver obtained from fish were cut in a Leica CM 3000 cryostat (Leica Instruments, Nusslock, Germany) at a cabinet temperature of −24 ◦ C and then stored (not longer than 24 h) at −40 ◦ C until staining. Prior to staining, slides were air-dried at 4 ◦ C for 20 min and another 10 min at room temperature. Then, sections were introduced into 0.1 M sodium citrate buffer (pH 4.5) containing 2.5% NaCl in intervals of 0, 2, 4, 6, 8, 10, 15, 20, 25, 30, 40 and 50 min in a shaking water bath at 37 ◦ C to destabilize the lysosomal membrane. Afterwards, sections were incubated for 10 min at 37 ◦ C in 0.1 M citrate buffer (pH 4.5) containing 2.5% NaCl, 0.04% naphthol AS-BI N-acetyl-␤-d-phosphate dissolved in 2% dimethyl sulfoxide and 7% of POLIPEP® as a tissue stabilizer. After incubation, sections were rinsed in 3% NaCl at 37 ◦ C for 5 min in a shaking water bath. Then, sections were transferred to 0.1 M phosphate buffer (pH 7.4) containing 0.1% of diazonium dye Fast Violet B salt for 10 min at room temperature. Slides were rinsed in running tap water for 10 min, fixed in 10% formaldehyde containing 2% calcium acetate for 15 min at 4 ◦ C and rinsed in distilled water. Finally, slides were mounted in Kaiser’s glycerin gelatin. Labilization period (LP) was determined under a Leitz Laborlux S light microscope (40× objective) as the maximum accumulation of reaction product associated with lysosomes. Four determinations were made for each animal by dividing each section into four approximately equal areas and assessing the first peak of labilization period in each area (Zorita et al., 2008). A mean value was then obtained for each section, corresponding to an individual liver. 2.7. Plasma vitellogenin (Vtg) levels Blood samples were obtained using heparinized syringes. After a brief centrifugation (500 × g for 5 min), plasma supernatant was quickly frozen in liquid nitrogen and stored at −80 ◦ C until further analysis. Plasma Vtg levels were measured by an indirect ELISA assay using a commercial polyclonal antibody against turbot Vtg (CS-2, Biosense Lab., Norway), according to the protocol described by Nilsen et al. (1998). Plasma samples were thawed on ice and diluted (1:50) in coating buffer (50 mM carbonate-bicarbonate buffer, pH 9.6). Diluted samples (100 ␮L) were added in triplicate to 96-well microtiter plates and incubated overnight at 4 ◦ C. Plates were rinsed three times with phosphate buffered saline solution containing 0.05% Tween 20 (PBST). Then,

wells were incubated for 1 h at room temperature with a blocking solution composed of PBST supplemented with 1% bovine serum albumin. After rinsing the plates with PBST, primary antibody diluted 1:1000 in blocking solution was added (100 ␮L) to each well and incubated for 2 h at room temperature. Plates were then rinsed with PBST and the secondary antibody (peroxidase conjugated goat anti-rabbit IgG, Sigma Chemical Co., St. Louis, MO, USA) diluted in PBST (1:10,000) was added (100 ␮L) to each well and incubated for 1 h at room temperature. After rinsing the plates with PBST, visualization of detected Vtg molecules was done by addition of 0.05 M phosphate–citrate buffer, pH 5 (100 ␮L) containing 0.012% hydrogen peroxide and 0.4 mg mL−1 o-phenylenediamine. Incubation was performed for 30 min at room temperature in the darkness. Reaction was stopped by adding 50 ␮L of 2 M H2 SO4 to each well. Absorbance was read at 492 nm. Non-specific binding was also measured for each plate, replacing samples by coating buffer alone. These wells were considered as blanks and average absorbance value of blanks from each plate was subtracted to each sample well.

2.8. Gonad and liver histopathology A portion of gonad tissue was fixed in 10% neutral buffered formalin and routinely processed for paraffin embedding in a Leica Tissue processor ASP 3000 (Leica Instruments, Nussloch, Germany). Sections of 5 ␮m thickness were cut in a Leitz 1512 microtome (Ernst Leitz, Vienna, Austria), stained with hematoxylin/eosin (Wilson and Gamble, 2002) and examined under a Leitz Laborlux S light microscope (Wetzlar, Germany) for histopathological alterations such as gamete abnormalities. Gametogenic developmental stages were determined for each animal, following the classification of Deng et al. (2007). A piece of the liver from each sampled fish was processed as above and prevalence of histopathological alterations, such as melanomacrophage centers (MMC), necrotic areas, vacuolization or parasites was determined. The samples were also examined to determine preneoplastic foci and neoplastic alterations according to the criteria established by Feist et al. (2004).

2.9. Statistical analyses Statistical analyses were carried out with the aid of the SPSS statistical package (V 14.0, SPSS Inc., Chicago, Illinois). For Vtg levels and AOX activity, differences along the exposure time were studied by one-way analysis of variance (ANOVA) followed by the Duncan’s test for multiple comparisons between pairs of means. Significant differences between control and exposed groups at each sampling time were studied using the Student’s t test. Previous to the analysis, data were tested for normality (Kolmogorov–Smirnov normality test) and homogeneity of variances (Levene’s test). In the case of LP and gene transcription levels, the non-parametric Kruskal–Wallis test was applied followed by the Mann–Whitney U test. For histopathological data, the Chisquare test was used. In all cases, significance was established at p < 0.05.

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0.80

Table 2 Concentration of 16 EPA PAHs in samples of heavy fuel oil no. 6. Results are given in ␮g g−1 . Mean ± SD

Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benzo[a]anthracene Chrysene Benzo[b + k]fluoranthene Benzo[a]pyrene Indeno[1,2,3-c,d]pyrene Dibenz[a,h]anthracene Benzo[g,h,i]perylene

686 ± 89 51 ± 3 272 ± 14 396 ± 20 1936 ± 116 213 ± 28 125 ± 14 516 ± 57 213 ± 13 464 ± 14 81 ± 11 167 ± 7 16 ± 3 27 ± 5 48 ± 3

PAHs

0 70 0.70 Styrene concen ntration (mg.L-1)

Compounds

121

0.60 0.50 0 40 0.40 0.30 0.20 0 10 0.10 0.00

5211

0

20

40

60

80 Time (h)

100

120

140

Fig. 2. Styrene concentration in seawater during the exposure period measured by SF-UV combined with GC–MS.

3. Results

3.3. Gene transcription levels

3.1. PAH concentrations in heavy fuel oil no. 6 and seawater during exposure

A fragment of 342 bp encoding ahr was sequenced in S. maximus and predicted AHR protein sequence showed 86% amino acid identity with AHR of Sebastiscus marmoratus (GenBank ID: ACH78368) (E value: 1e−51 ). For ppar˛, 200 bp were amplified from turbot liver. The predicted protein sequence showed 95% amino acid identity with Sparus arurata PPAR␣ (GenBank ID: ABB29464) (E value: 2e−37 ). Obtained sequences for turbot ahr and ppar˛, together with published sequences of cyp1a1 and EF1-˛, were used to measure transcription levels of those genes by quantitative real-time RT-PCR. cyp1a1 transcription level was significantly increased in liver after 3 and 14 days of exposure to heavy fuel oil. After the recovery period, transcription levels returned to control levels (Fig. 3a). ahr transcription level did not show significant variations between oiltreated and control turbots along the experiment, however higher transcription levels were measured in treated turbots than in control animals after 14 days of exposure and 14 days in clean water (Fig. 3b). Transcription level of ppar˛ did not vary significantly between oil-treated and control animals along the experiment, but after the recovery period, transcription was significantly elevated in turbots previously exposed to heavy fuel oil in comparison to the control group (Fig. 3c).

The results of the PAH analyses in the heavy fuel oil no. 6 sample are shown in Table 2. Concentration ranged from 16 ␮g g−1 of indeno[1,2,3-c,d]pyrene to 1936 ␮g g−1 of phenanthrene. The sum of the 16 PAHs reached 5211 ␮g g−1 . The PAH concentration in seawater along the experiment ranged from 311.43 to 4386.90 ng L−1 (Table 3). In the exposure tank, the highest PAH concentration was measured before fish were introduced in the tank (4386.90 ng L−1 ). However, the PAH concentration decreased considerably at day 2 (483.98 ng L−1 ).

3.2. Water and tissue concentrations of styrene SF-UV combined with GC–MS analysis showed that styrene concentration in seawater fluctuated along the exposure period reaching up to 0.7 mg L−1 (Fig. 2). Styrene concentration in the tissues of exposed fish was about 18.75 ± 12.03 ␮g g−1 whereas styrene concentration in control fish was close to the GC–MS detection limit.

Table 3 Concentration of 16 EPA PAHs in seawater from the exposure tank along the experiment. Results are given in ng L−1 . bdl: below detection limit. Compounds

Outflow

T0

1 day

2 days

3 days

7 days

14 days

Naphthalene Acenaphthylene Acenaphthene Fluorene Phenanthrene Anthracene Fluoranthene Pyrene Benzo[a]anthracene Chrysene Benzo[b + k]fluoranthene Benzo[a]pyrene Indeno[1,2,3-c,d]pyrene Dibenz[a,h]anthracene Benzo[g,h,i]perylene

52289.91 26.46 4258.46 4463.70 708.43 62.61 6.86 36.80 229.05 258.79 bdl bdl bdl bdl bdl

3499.63 1.47 225.79 203.76 356.66 56.21 5.57 18.15 9.06 10.60 bdl bdl bdl bdl bdl

1885.35 48.27 194.61 138.49 142.27 57.22 0.00 16.40 121.78 38.90 bdl bdl bdl bdl bdl

99.11 17.78 215.27 34.57 30.36 35.82 8.81 29.73 4.18 8.35 bdl bdl bdl bdl bdl

71.53 55.97 98.68 50.68 13.89 28.39 3.94 16.81 2.16 6.21 bdl bdl bdl bdl bdl

26.90 62.40 31.42 53.43 17.65 20.25 19.10 20.80 2.57 6.36 bdl bdl bdl bdl bdl

9.81 40.00 78.38 54.94 32.55 40.20 5.44 36.94 4.90 8.27 bdl bdl bdl bdl bdl

PAHs

62341.06

4386.90

2643.29

483.98

348.26

260.86

311.43

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cyp1a1

cyp1a1

3

*

30 Q RQ

d

40

Styrene

2

*

20

RQ

a

Fuel oil

1 10 0

0 3 days

b

14 days

3 days

Recovery (14 days)

ahr

e

30

6

7 days

Recovery (7 days)

7 days

Recovery (7 days)

7 days

Recovery (7 days)

ahr

25 4

RQ

RQ

20 15

2

10 5

0

0 3 days

c

14 days

3 days

Recovery (14 days)

pparα

f 15

30

pparα

25 10 RQ

RQ

20 15

*

10

5

*

5 0

0 3 days

14 days

Recovery (14 days)

Control Heavy fuel oil

3 days

Control Styrene

Fig. 3. Transcription levels of cyp1a, ahr and ppar˛ normalized to EF1-˛ determined in liver of turbots (n = 8) exposed to heavy fuel oil (a–c), and to styrene (d–f). Values are given as means and standard deviations. Asterisks indicate significant differences between control and exposed turbots according to the Mann–Whitney U test, p < 0.05. RQ, relative quantification.

On the other hand, exposure to styrene did not affect transcription of cyp1a1 and ahr genes (Fig. 3d and e), but up-regulated significantly ppar˛ transcription after 3 days of exposure (Fig. 3f). 3.4. Acyl-CoA oxidase activity Compared to the control group, liver peroxisomal AOX activity increased significantly after 14 days of exposure to fuel oil. After the recovery period, AOX activity decreased significantly compared to the control group and to the oil-exposed group at 14 days of exposure (Fig. 4a). In fish exposed to styrene, AOX activity was significantly inhibited compared to controls only in the group sampled after the recovery period (Fig. 4b).

3.5. Lysosomal membrane stability In general, LP values were low in all fish from both heavy fuel oil and styrene experiments. Treated fish presented significantly lower LP values than in the control group after 14 days of heavy fuel oil exposure (Fig. 5a) and 7 days of styrene exposure (Fig. 5b). After the recovery period, LP values in both treated groups of turbots increased again up to control values. 3.6. Plasma vitellogenin levels Levels of Vtg in the plasma of experimental fish were under detection limit in most cases (data not shown). In the heavy fuel

100 90.90 90 100 83.33 63.63 100 100 0 18.18 0 8.33 8.33 9.09 8.33 20 18.18 9.09 20 8.33 16.66 9.09 8.33 20 18.18 9.09 10 0 16.66 27.27 16.66 0 18.18 36.36 50 58.33 8.33 18.18 33.33 50 45.45 36.36 30 33.33 41.66 45.45 50 40 MMC: melanomacrophage center.

81.81 63.63 90 75 58.33 81.81 75 50 T0 3 days 7 days Recovery (7 days)

Styrene Control Styrene Control Styrene Control Styrene Control Styrene Control Styrene Styrene Control

Control

41.66 27.27 75 60

Oil Control

Parasites

27.27 27.27 75 75 8.33 18.18 41.66* 60* 27.77 18.18 0 10 8.33 9.69 16.66 0 0 0 8.33 30 8.33 0 8.33 20 0 0 0 10 16.66 9.09 16.66 60 27.27 0 33.33 60 41.66 54.54 41.66 20 36.36 36.36 16.66 10 50 18.18 25* 60

Oil

Extensive cell vacuolization

Control Oil

Focal cell vacuolization

Control Oil

Extensive necrosis

Control Oil Control

Areas of necrosis

Oil Control

27.27 9.09 66.66 60 T0 7 days 14 days Recovery (14 days)

Fig. 5. Labilization period of lysosomal membrane in livers of control and exposed fish from the oil (a) and styrene (b) experiments. Values are given as means and standard deviations. Asterisks indicate significant differences with respect to controls of the same day according to the Mann–Whitney U test at p < 0.05.

Focal necrosis

Both control and treated female turbots showed gonad at stage 1 of gametogenesis (gonad composed of primary and secondary oocytes) during the whole experiment, as corresponds to juvenile animals. In male turbots, stages 1 (only spermatogonia) and 2 (spermatogonia and spermatocytes) were observed. No differences in gamete development were detected between control and oil- or

Oil

3.7. Gonad and liver histopathology

Control

oil experiment, only some animals in the control group at T0 showed detectable Vtg levels in plasma. In the styrene experiment, some few females distributed in the different experimental groups showed detectable Vtg levels.

123

MMC

Fig. 4. AOX activity in the livers of control and exposed fish from oil (a) and styrene (b) experiments. Values are given as means and standard deviations. Asterisks indicate significant differences with respect to controls of the same day according to the Student’s t test at p < 0.05.

Table 4 Prevalence of histopathological alterations in livers of turbots. Data are shown in percentages. Asterisks indicate statistically significant differences between control and exposed groups (p < 0.05) according to Chi-square test.

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Fig. 6. Micrographs of the liver of turbots exposed to oil and styrene showing various pathologies. (A) Extensive cell vacuolization in liver of a turbot exposed to oil for 3 days. (B) Encapsulated parasites (arrows) in the liver of a turbot exposed to oil for 14 days. (C) MMC (asterisks) in liver of a turbot exposed to styrene for 7 days. (D) Extensive necrosis (arrows) in a turbot sampled 3 days after exposure to styrene. Bars: 100 ␮m (a and c), 200 ␮m (b and d).

styrene-exposed fish. Histopathological alterations were not found in the gonad tissue of experimental animals. The prevalence of the histopathological alterations found in the liver of turbots is shown in Table 4. Livers with accumulation of MMC and focal necrosis were common in most groups, but prevalence of MMC was significantly lower in turbots exposed to oil for 14 days than in control fish sampled at the same time. Also in this sampling, the prevalence of extensive cell vacuolization (Fig. 6a) increased significantly in oil-exposed fish compared to control fish and no recovery was observed (Table 4). Parasites (Fig. 6b) were common in both experimental groups, showing no differences between them. In the styrene experiment, no significant differences were found between treated and control animals for any of the pathologies studied, although for some of them, such as MMC (Fig. 6c) and necrotic areas (Fig. 6d), styrene-exposed animals showed higher prevalences in most samplings. 4. Discussion This study aimed to determine the short-term effects of heavy fuel oil no. 6 and styrene, using a battery of molecular, cell and tissue-level biomarkers in juvenile turbots. The possible recovery of biomarker responses was also evaluated by placing fish in clean water after exposure. The concentration of measured PAHs in the samples of the heavy fuel oil no. 6 was 5211 ␮g g−1 . This value is comparable to that reported for the Prestige heavy fuel oil, where

concentration of the 16 EPA PAHs was 1210.7 ␮g g−1 (Alzaga et al., 2004). Accordingly, in North Sea crude oil the concentration of the 16 EPA PAHs was 1782 ␮g g−1 (Aas et al., 2000). The resulting PAH concentrations in the water column measured in this study (311.43–4386.9 ng L−1 ) were similar to those described after different oil spills. For instance, 1 month after the Prestige oil spill, González et al. (2006) reported seawater PAH concentrations up to 2090 ng L−1 . Nevertheless, a wide range of PAH concentrations in seawater are described in the literature after oil spills: PAH concentration reached values from 20.9 to 6000 ng L−1 after the Erika and Exxon Valdez oil spills, respectively (Short and Harris, 1996; ´ Tronczynski et al., 2004). At the beginning of the experiment, before fish were introduced in the tank, the PAH concentration in seawater was 4386.9 ng L−1 . After two days of exposure, PAH concentration in seawater decreased approximately ten times (483.98 ng L−1 ). This decline of PAH concentration is explained by the volatilization of low molecular weight PAHs, such as naphthalene, as well as by the introduction of fish that are able to take up PAHs. In a similar study where mullets were exposed to fresh heavy fuel oil, the PAH concentration in seawater decreased from 97,180 at 2 days of exposure to 3500 ng L−1 at day 16 (Bilbao et al., 2010). Even though the tanks were closed, styrene concentration in seawater fluctuated throughout the exposure, reaching levels of up to 0.7 mg L−1 . Information on styrene levels in the aquatic environment is limited. Styrene concentrations in water are usually less than 20 ␮g L−1 (Alexander, 1997). Nevertheless, concentrations up

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to 100 ␮g L−1 have been detected in a small number of industrial effluents in the US (Alexander, 1990). Information on the toxicity of styrene to aquatic organisms is restricted to acute studies, as styrene is a highly volatile compound and its half-life in water is only about 3 h (Fu and Alexander, 1992). In fathead minnows, the LC50 and NOEC values for styrene were 10 and 4 mg L−1 at 96 h, respectively, indicating a moderate toxicity for these animals (Cushman et al., 1997). Mamaca et al. (2005) showed that styrene concentration in seawater decreased from the nominal concentration of 2–0.2 mg L−1 during Symphodus melops exposure. Volatilization and biotransformation are the main processes accounting for the loss of styrene when it reaches the environment (Fu and Alexander, 1992; Gibbs and Mulligan, 1997). In spite of the low styrene concentration in seawater the results of this study indicate the high potential of turbots to accumulate styrene. As fish are able to rapidly metabolize hydrocarbons in the liver, styrene accumulation was measured in the muscle, where detected concentration was high (18.75 ␮g g−1 ) when compared with a previous study in which styrene concentration in crab tissues was 0.093 ␮g g−1 7 days after the Ievoli Sun spill (Law et al., 2003). Biotransformation is among the most studied processes in aquatic animals exposed to organic xenobiotics. Transcriptional induction of cyp1a1 in fish may serve as a sensitive marker of xenobiotic exposure and early biological response (Hahn and Stegeman, 1994). Induction of cyp1a1 transcription has been observed in fish following laboratory exposure to PAH compounds, sediment extracts, and also in fish captured from PAH-contaminated environments (Bilbao et al., 2010; Courtenay et al., 1999; Meucci and Arukwe, 2006; Paetzold et al., 2009). In the present study, cyp1a1 transcription was significantly induced in liver after 3 and 14 days of exposure to heavy fuel oil. This response has also been observed in mullets after exposure to Prestige oil (Bilbao et al., 2010). After 14 days in clean water, cyp1a1 transcription level returned to control levels, which indicates the reversibility of the response after inducing compounds have been removed from water. In mammals, styrene degradation is also catalyzed by cytochrome P450-dependent monoxygenases (Gibbs and Mulligan, 1997; Rueff et al., 2009). However, in this work no changes were observed in cyp1a1 liver transcription levels after exposure to styrene. It is well documented that cyp1a1 transcription in response to PAH exposure is mediated through activation of AhR pathway (Zhou et al., 2005) and heavy fuel oil is an AhR agonist (Hahn, 2002). Different to mammals, fish have two isoforms of the ahr gene (ahr1 and ahr2) (Andreasen et al., 2002; Hahn et al., 1997; Hansson et al., 2003). Although the respective functions of these two ahr forms are not well understood, ahr2 appears to play a major role in mediating the toxicity of PAHs (Clark et al., 2010). The ahr sequence obtained in turbot was more similar to ahr1 than ahr2 from different fish species. In our study, heavy fuel oil and styrene exposure did not produce any significant change in ahr transcription levels at the measured timepoints, although exposed turbots presented higher levels than control animals. PAHs have been described to produce peroxisome proliferation in several aquatic organisms (Cajaraville et al., 2003; Oakes et al., 2005; Ortiz-Zarragoitia and Cajaraville, 2005; Ortiz-Zarragoitia et al., 2006). Similar to mammals, peroxisome proliferators act in fish by binding to PPARs (Colliar et al., 2011). PPARs are members of the nuclear receptor superfamily of ligand-activated transcription factors (Mandard et al., 2004). Target genes, like those of the peroxisomal ␤-oxidation pathway in the case of ppar˛, are mainly involved in lipid homeostasis (Mandard et al., 2004). Turbots exposed to heavy fuel oil showed higher levels of hepatic transcription of ppar␣ compared with controls, but this difference was statistically significant only after 14 days in clean water. This could indicate that previous oil exposure can affect ppar˛

125

transcription and that effects could be observed even when the contaminant was removed. In case of styrene, ppar˛ transcription level was significantly up-regulated after 3 days of exposure, suggesting that the response might appear more rapidly after exposure to styrene than to heavy fuel oil. One of the genes whose transcription is regulated by ppar˛ is AOX (Dreyer et al., 1992), the first and rate limiting enzyme of the peroxisomal fatty acid ␤-oxidation pathway. AOX is generally used as a marker of peroxisome proliferation since its activity is induced when sensitive organisms are exposed to peroxisome proliferators (Cajaraville et al., 2003). Thus, induction of AOX activity has been measured as early indicator of exposure to organic xenobiotics (Bilbao et al., 2006; Cajaraville et al., 2003; Fahimi and Cajaraville, 1995; Gunawickrama et al., 2008; Holth et al., 2011; Zorita et al., 2008). In this study, AOX activity was significantly increased after 14 days of exposure to heavy fuel oil in comparison with control values. In grey mullets (Chelon labrosus) exposed to Prestige-like fuel oil, AOX activity was induced after 2 and 16 days of exposure (Bilbao et al., 2010). In contrast, we did not find differences in AOX activity in animals exposed for up to 7 days to styrene, suggesting that styrene did not act as a peroxisome proliferator in turbot liver under the assayed conditions. Thus, the significant up-regulation of ppar˛ transcription after 3 days of styrene exposure, was not followed by an induction in AOX activity. Similarly, in mullets, up-regulation of ppar˛ transcription after 2 days of exposure to weathered fuel oil was not immediately reflected in induced AOX activity. Increased AOX activity was not recorded until day 16, suggesting that observed regulation in transcription levels of the ppar˛ receptor needed more time to elicit a response at protein activity level (Bilbao et al., 2010). Lysosomes have a crucial role in the detoxification of toxic substances and constitute an important target of toxicants effects. For these reasons, changes in the lysosomal system have been used as general marker of pollutant impact in a number of field studies using fish as sentinel organisms (Köhler et al., 2002; Zorita et al., 2008). Both control and exposed turbot of this study presented LP values below 10 min, which indicate a bad condition of fish according to established baseline and critical values for environmental health assessment (Köhler et al., 2002). Nevertheless, low LP values (6–7 min) have been reported in eelpout and in flounder collected from reference sites in the Baltic Sea (Barˇsiene et al., 2006; Lehtonen et al., 2006). The range of LP values reported in control fish could be attributed to species differences or to differences in sample preparation between laboratories. Even so, lysosomal membrane stability was significantly reduced in hepatocytes of oiland styrene-exposed turbots in comparison with control turbots after 14 days and 7 days of exposure, respectively. This indicates that tested contaminants were taken up and produced deleterious effects in treated fish. After the recovery period in clean water, LP values of oil- and styrene-exposed turbots increased up to control values, suggesting that the depuration period in both experiments was enough to enable them to recover the lysosomal membrane stability up to control values. In agreement with our results, thicklip grey mullets exposed under laboratory conditions to Prestige-like fresh and weathered heavy fuel oil for 2 and 16 days showed significantly lower LP values than the controls (Bilbao et al., 2010). Thus, lysosomal membrane stability has to be considered as an indicator of non-specific general physiological stress and may be used to assess the recovery of health impairment produced by fuel oil and styrene. Vtg levels have been widely used as a biomarker of exposure to xenoestrogens in fish (Arukwe and Goksøyr, 2003). Levels of Vtg in plasma of experimental fish were under detection limit in most cases, suggesting that heavy fuel oil and styrene did not act as typical xenoestrogens in turbots. Zebrafish exposed to the water accommodated fraction of crude oil showed antiestrogenic

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effects (Arukwe et al., 2008). The antiestrogenicity of PAH (3methylcholanthrene) has also been reported in rainbow trout hepatocytes in vitro (Navas and Segner, 2000). Similarly, immature rainbow trout exposed to naphthalene showed reduced plasma 17␤-estradiol levels and reduced reserves directed to vitellogenesis and gamete development (Tintos et al., 2006). In vitro ER binding assays showed a low affinity of styrene for the estrogen receptor. Besides, more specific in vitro reporter gene assays and in vivo rat uterotrophic assay did not show estrogenic effects caused by styrene oligomer exposure (Ohno et al., 2003). Accordingly, control and exposed turbots in both experiments showed gametes at early stages of development (stage 1) during the whole experiment. No differences in gamete development existed between control and exposed fish, indicating that reproductive potential was not disturbed by heavy fuel oil or styrene exposure. Strong evidence for a relationship between water pollution and fish disease has been found by several researchers (Grinwis et al., 2000; Lang et al., 2006; Stehr et al., 1998). The occurrence of liver lesions considered to be associated with exposure to anthropogenic contaminants has been recorded in fish from several coastal areas (Lang et al., 2006; Myers et al., 1998; Stentiford et al., 2003; Vethaak and Jol, 1996). Laboratory experiments with PAHs performed with several fish species have also resulted in different histopathological alterations (Bilbao et al., 2010; Reynolds et al., 2003). In our study, control animals of both experiments presented several histopathological alterations including accumulation of MMC, necrosis, vacuolization and presence of parasites. These alterations were also observed at the beginning of the experiments, which could be related to the low LP values observed. A significant increase of extensive cell vacuolization was observed at 14 days and after recovery period in the liver of turbots exposed to fuel oil. This finding is similar to those reported for flounders caught from polluted UK estuaries (Lyon et al., 2004) and for salmons after 10 days of exposure to the water soluble fraction of Alaska North Slope crude oil (Brand et al., 2001). Therefore, it could be suggested that fuel oil exposure caused effects in fish liver at short time that remained after 14 days in clean water. Prevalence of MMC in turbots decreased significantly after 14 days of fuel oil exposure, in agreement with previous studies on fuel oil effects in mullets (Bilbao et al., 2010). In flounders, chronic exposure to PAH also caused a significant decreased in MMC amount (Payne and Fancey, 1989) and reduced activity of melanomacrophage aggregates was observed in individuals from polluted areas (Broeg, 2003), indicating severe effects of PAHs on the immune system of chronically exposed fish. No significant liver alterations were observed in turbots exposed to styrene although a trend for a higher prevalence of MMC, areas of necrosis and focal cell vacuolization was identified in styreneexposed turbots compared to control animals. Recovery was only observed in focal cell vacuolization after 7 days in clean water. As a summary, transcript of the PAH-responsive gene cyp1a1 was significantly overexpressed after exposure to heavy fuel oil. Peroxisomal AOX activity was induced 14 days after exposure. Both responses disappeared after withdrawal of the contaminant. Lysosomal membrane stability test gave consistent results with reduced LP after 14 days of exposure to heavy fuel oil and further recovery. As indicated by Vtg data, heavy fuel oil exposure did not produce xenoestrogenic effects in turbots. Histopathological examination of gonad indicated that heavy fuel oil did not provoke effects on gamete development. However, significant histopathological alterations were observed in liver of oiled turbots. Styrene did not induce alterations in transcription levels of genes related to biotransformation whereas ppar˛ transcription level was up-regulated after 3 days exposure. However, styrene did not behave as a typical AOX inducing peroxisome proliferator in turbot. Styrene provoked a general stress: however, we did not observe evidences of effects on gamete development and liver histology. Finally, according to

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