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Aquatic Toxicology 116– 117 (2012) 116– 128

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Aquatic Toxicology

j ourna l ho me p ag e: www.elsev ier .com/ l ocate /aquatox

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

Pamela Ruiza, Maren Ortiz-Zarragoitiaa, Amaia Orbeaa, Michael Theronb, Stéphane Le Flochc,Miren P. Cajaravillea,∗

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, Spainb Laboratoire ORPHY, EA 4324, UFR Sciences et Techniques, Université de Bretagne Occidentale, 6 avenue Le Gorgeu, CS 93837, 29238 Brest Cedex 3, Francec Centre of Documentation, Research and Experimentations on Accidental Water Pollution, 715 rue Alain Colas, CS 41836, 29218 Brest Cedex 2, France

a r t i c l e i n f o

Article history:Received 3 October 2011Received in revised form 17 January 2012Accepted 5 February 2012

Keywords:Conventional and molecular biomarkersHeavy fuel oil no. 6StyreneRecoveryTurbots

a b s t r a c t

Several accidental spills in European coastal areas have resulted in the release of different toxic com-pounds into the marine environment, such as heavy fuel oil type no. 6 in the “Erika” and “Prestige” oilspills and the highly toxic styrene after the loss of the “Ievoli Sun”. There is a clear need to develop toolsthat might allow assessing the biological impact of these accidental spills on aquatic organisms. The aimof the present study was to determine the short-term effects and recovery after exposure of juvenilefish (Scophthalmus maximus) to heavy fuel oil no. 6 and styrene by using a battery of molecular, cell andtissue level biomarkers. Turbots were exposed to styrene for 7 days and to the diluted soluble fractionof 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, whereasahr transcription was not modulated after heavy fuel oil and styrene exposure. ppar˛ transcription levelwas significantly up-regulated after 3 days of treatment with styrene. Liver activity of peroxisomal acyl-CoA oxidase (AOX) was significantly induced after 14 days of oil exposure, but it was not affected bystyrene. Hepatocyte lysosomal membrane stability (LMS) was significantly reduced after exposure toboth treatments, indicating that the tested compounds significantly impaired fish health. Both AOX andLMS values returned to control levels after the recovery period. No differences in gamete developmentwere observed between fuel- or styrene- exposed fish and control fish, and vitellogenin plasma levelswere low, suggesting no xenoestrogenic effects of fuel oil or styrene. While styrene did not cause anyincrease in the prevalence of liver histopathological alterations, prevalence of extensive cell vacuoliza-tion increased after exposure to heavy fuel oil for 14 days. In conclusion, the suite of selected biomarkersproved to be useful to determine the early impact of and recovery from exposure to tested compoundsin turbot.

© 2012 Published by Elsevier B.V.

1. Introduction

Accidental spills in the marine environment cause significantchemical pollution, with acute effects on exposed organisms (Kirbyand Law, 2010). A large number of marine ecosystems have beenaffected by spills. The “Exxon Valdez” oil spill caused the release of42,000 tons of crude oil in Alaska in 1989 (Spies et al., 1996); the

Abbreviations: AHR, aryl hydrocarbon receptor; AOX, peroxisomal acyl-CoAoxidase; 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 aromatichydrocarbon; PPAR, peroxisome proliferator-activated receptor; RQ, relative quan-tification; VTG, vitellogenin; WSF, water soluble fraction.

∗ Corresponding author. Tel.: +34 94 6012697; fax: +34 94 6013500.E-mail address: [email protected] (M.P. Cajaraville).

tanker “Erika” lost 20,000 tons of fuel oil (classified as no. 2 accord-ing to Association Franc aise de Normalisation and no. 6 accordingto the American Society for Testing and Materials) in front of thecoast of Brittany in 1999 (Bocquené et al., 2004); the sinking of thetanker “Ievoli Sun” resulted in the release of more than 1000 tonsof the hydrocarbon styrene in the English Channel in 2000 (Lawet al., 2003); the tanker “Prestige” sunk in front of the Galician coastin 2002 spilling more than 60,000 tons of heavy fuel oil (Gonzálezet al., 2006), and more recently the spill which occurred in theGulf of Mexico released 780,000 tons of oil to the sea impactingthe Southeast coastline of USA (Mitsch, 2010).

The spilled compounds may affect the organisms producingchanges at molecular, cellular and physiological levels whichat long-term may provoke the decline of their populations, asshown after the “Exxon Valdez” oil spill (Peterson, 2001). Severalworks have employed biomarkers to assess biological effects after

0166-445X/$ – see front matter © 2012 Published by Elsevier B.V.doi:10.1016/j.aquatox.2012.02.004


P. Ruiz et al. / Aquatic Toxicology 116– 117 (2012) 116– 128 117

accidental spills. Ten years after the “Exxon Valdez” oil spill, expo-sure of fish to oil was still detected by means of ethoxyresorufinO-deethylase (EROD) activity and biliary fluorescent metabolites(Huggett et al., 2003; Jewett et al., 2002). Likewise, 1 year afterthe “Prestige” oil spill peroxisomal acyl-CoA oxidase (AOX) activity,lysosomal responses such as changes in the structure and mem-brane stability and histopathology in mussel and fish discriminatedthe most impacted areas by the spilled oil (Marigómez et al., 2006;Orbea et al., 2006). Laboratory studies have also been carried outto assess the biological effects of polycyclic aromatic hydrocarbons(PAHs) in aquatic animals. PAHs are the main components of thesespilled oils (Alzaga et al., 2004) and are able to induce biotransfor-mation metabolism and peroxisome proliferation in marine organ-isms (Cajaraville et al., 2003; Bilbao et al., 2010). In vertebrates,biotransformation and peroxisome proliferation are mediated byaryl hydrocarbon receptor (AhR) and by peroxisome proliferator-activated receptors (PPARs) (Barron et al., 2004; Mandard et al.,2004), respectively. PAHs are also the oil components that pose thehighest environmental risk, mainly due to their carcinogenic andmutagenic properties (Beyer et al., 2010). PAHs are absorbed byfish via the gills and body surface, but also by ingestion of food orthrough contaminated sediment (Van der Oost et al., 2003).

Compared to other aromatic compounds, little attention hasbeen paid to styrene, probably due to its lower rate of industrialuse in comparison with PAHs (Cushman et al., 1997; Gibbs andMulligan, 1997; Mamaca et al., 2005). Styrene is primarily used inthe production of polymers and polystyrene. Styrene is commonlyshipped by vessel and due to its characteristics (volatility and lowsolubility in water) it has rarely been detected in water and, whenpresent, it occurs at very low levels (ng L−1) (Cushman et al., 1997).Based on the results of toxicological studies and on measured con-centrations in water, styrene is not deemed to cause effects onaquatic 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 determinethe short-term effects of and recovery from laboratory exposureof juvenile turbots (Scophthalmus maximus) to two compoundsappearing in the environment as a result of accidental spills, thecomplex mixture formed by heavy fuel oil no. 6, rich in PAHs, andthe highly volatile hydrocarbon, styrene. The experimental systemwas designed to simulate a realistic scenario immediately aftera spill. Juvenile turbots were exposed to the water soluble frac-tion of fuel oil no. 6 for 14 days and to styrene for 7 days, andthen transferred to clean seawater for the same period of time. Alonger exposure period for styrene would not be realistic or rele-vant, because styrene evaporates quickly in open systems and, thus,high concentrations of this compound would not last in the aquaticenvironment (Alexander, 1997). On the other hand, oil componentsare 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 mon-itored in seawater and, at the end of the experiment, styreneconcentration in tissues was quantified to evaluate bioaccumula-tion processes. An integrated battery of biomarkers from molecularto cell and tissue levels were measured which reflects impact atdifferent levels of biological organization. Peroxisome prolifera-tion and induction of biotransformation metabolism were studiedas exposure biomarkers for organic pollutants. Peroxisome pro-liferation was assessed at gene transcription level (ppar˛) and atAOX activity level (Cajaraville et al., 2003; Zorita et al., 2008).Induction of biotransformation metabolism was studied by ahrand cyp1a1 transcription level (Hahn, 2002; Zhou et al., 2005).Lysosomal membrane destabilization in fish hepatocytes was usedas an indicator of non-specific effects arising from an increasedintracellular accumulation of xenobiotics (Köhler, 1991). Plasma

vitellogenin levels were measured as biomarker of estrogenicity(Arukwe and Goksøyr, 2003) and gonad histology was also eval-uated as supporting parameter. Finally, livers were examinatedfor histopathological alterations since several studies carried outin coastal waters have shown correlation between environmen-tal contaminants and the occurrence of toxicopathic liver lesionsin fish (Feist et al., 2004; Stentiford et al., 2003; Vethaak et al.,1996). In recent years, fish diseases and liver histopathologicalalterations have been used as indicators of pollution effects andhave been implemented in monitoring programs (Feist et al., 2004;Lang, 2002). The presence of inflammatory lesions, hepatocellu-lar fibrillar inclusions, and preneoplastic and neoplastic lesions ishigher in fish captured in polluted environments than in fish fromreference sites (Stentiford et al., 2003).

2. Materials and methods

2.1. Animals

Juvenile turbots, S. maximus (n = 500; weight 358 ± 93 g, length28 ± 2 cm, mean ± SD) were purchased from a fish farm (France Tur-bot, Tredarzec, France). Before starting the experiment, they wereacclimatized for 20 days in two laboratory tanks of 1000 L (controltank and treatment tank). They were fed daily with dried pellets(Aquaculture food Le Gouessant®, Lamballe, France, 4.5 mm diam-eter, 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 bothtanks. Water was taken from Sea of Iroise, France, and sterilized bymeans of UV-rays and filters system. Experiments were performedin the facilities of Cedre (Brest, France).

2.2. Experimental systems

2.2.1. Exposure of turbots to fuel oil number 6Experiments were carried out in November 2006. The light

regime was set according to the season: 14 h light, 10 h dark. Watersalinity (35–36‰), water pH (7.98 ± 0.06), oxygen concentration(228 ± 18 �mol L−1) and sea water temperature (16 ± 2 ◦C) weremeasured daily. During the experiment, turbots were fed twice perday with dried pellets.

Turbots were exposed for 14 days to 10% of the water solublefraction (WSF) of heavy fuel oil no. 6 using a continuous flow-through system with water aeration described by Aas et al. (2000).The system made it possible to expose fish to relative stable con-centrations of dispersed crude oil (Aas et al., 2000), similar to thosereported after accidental oil spills such as Prestige and Exxon Valdezshipwrecks: 2090 ng L−1 (González et al., 2006) and 6000 ng L−1

(Short and Harris, 1996), respectively. The experimental systemconsisted of a 2 m high column, which was filled with glass beadscoated with oil and seawater was running through the column fromtop to bottom, to produce the WSF of the oil, which was diluted tentimes before being administered to the animals. After the expo-sure period, fish were maintained for 14 additional days in cleanseawater for recovery.

A sample of heavy fuel oil was collected to characterize the PAHcomposition of fuel oil used in the experiment. Seawater sampleswere taken at T0 (before fish were introduced), and after 1, 2, 3, 7and 14 days of exposure to monitor PAH concentration in the watercolumn of the contamination tank along the exposure period.

Fish samples were collected before exposure started (T0) andat 3, 7 and 14 days of exposure. A last sampling was done afteradditional 14 days of recovery in clean seawater.

At each sampling time of the exposure period, 12 fish per exper-imental group were sacrificed. From these samples, a piece of liverof 7–8 turbots per experimental group was immersed individually


118 P. Ruiz et al. / Aquatic Toxicology 116– 117 (2012) 116– 128

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

in RNA later® (Sigma–Aldrich, St. Louis, Missouri, USA) and frozenin liquid nitrogen for transcription studies. Livers, gonads and bloodsamples of 12 turbots per experimental group were collected andprocessed individually at each sampling time for AOX activity,plasma vitellogenin level, and gonad and liver histopathologicalanalysis. For lysosomal membrane stability test, a small piece ofthe liver obtained from 5 fish per experimental group were rapidlyfrozen in liquid nitrogen and stored at −80 ◦C until sectioning.

After the recovery period, 5 fish previously exposed to fuel oiland 12 control fish were sacrificed and samples for the all abovedeterminations were collected from each individual.

2.2.2. Exposure of turbots to styreneTurbots 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, simulatingthe possible concentration present in the sea surface after a styrenespill. The experimental system was composed of a mixing tank con-nected to an exposure tank and a degassing column (Fig. 1). In themixing tank, which had a capacity of 316 L, a slick of styrene wasplaced in contact with the aqueous phase and dissolved slowly.Then, contaminated water containing the dissolved fraction of theslick was pumped to the exposure tank (300 L) where turbots weremaintained. Styrene disappeared rapidly from water and, there-fore, the concentration remaining in the exposure tank would besimilar to that appearing in seawater immediately after a spill.

The exposure tank was connected to a system of waste waterrecuperation by overflow. During the experimentation, a waterflow of 1.3 L min−1 was pumped and sent on a degassing col-umn before being directed to the mixing tank. A second circuitensured a renewal of fresh sea water to maintain the nitrate andnitrite concentrations stable. In addition, the oxygen saturationof water was maintained into this tank by a compressor, whichinjected air via diffuser. The degassing column favored the gaseous

exchange, allowed water oxygenation by percolation on glass beadsand allowed elimination of dissolved carbon dioxide produced byorganisms (carbon dioxide transfer from water to air). During theexperimentation, both mixing and exposure tanks were kept cov-ered to avoid cross contamination from the styrene exposure tankto the control tank. The concentration of styrene was monitored inseawater during the experiment.

Fish samples were collected before the exposure (T0), after 3and 7 days of exposure and after 7 days of recovery in clean water.

Muscle tissues of 10 control and 10 styrene-exposed turbotswere collected after 7 days of exposure to determine the levelof styrene in fish. Styrene concentration was measured in mus-cle because xenobiotic metabolism in this organ is much slowerthan in liver and, thus, styrene accumulation in muscle could betterreflect environmental exposure. Livers of 6-7 turbots per experi-mental group at each sampling time were dissected out, immersedindividually in RNA later® and frozen in liquid nitrogen for tran-scription studies. Livers, gonads and blood samples of 12 turbotsper experimental group were collected and processed individuallyat each sampling time for liver AOX activity, plasma vitellogeninlevels and histopathological analysis of liver and gonad. For lyso-somal membrane stability test a small piece of the liver obtainedfrom 5 fish per experimental group were rapidly frozen in liquidnitrogen and stored at −80 ◦C until sectioning.

2.3. Chemical analyses

2.3.1. PAH concentrations in heavy fuel oil no. 6 and seawaterduring exposure

The composition of the oil used in this study was deter-mined according to Wang and Fingas (1995). The 16 priorityPAHs of the U.S. Environmental Protection Agency (EPA)were measured (naphthalene, acenaphthylene, acenaph-thene, fluorene, phenanthrene, anthracene, fluoranthene,pyrene, benzo[a]anthracene, chrysene, benzo[b]fluoranthene,



benzo[k]fluoranthene, benzo[a]pyrene, indeno[1,2,3-c,d]pyrene,dibenz[a,h]anthracene and benzo[g,h,i]perylene). Around20–30 mg of the heavy fuel oil was spiked with 100 �L of amixture of internal standards (perdeuterated PAHs) and puri-fied by solid phase extraction (cyanopropylphase, compoundseluted by 5 mL of a mixture pentane/dichloromethane 80/20).The sample volume was then reduced to 200 �L prior to thegas chromatography coupled to mass spectrometry (GC–MS)analysis. The extracted compounds were analyzed on a HP 5890series II gas chromatograph coupled to a HP 5973 mass selec-tive detector (MSD) (Electronic impact: 70 eV, voltage: 1300 V)(Hewlett-Packard, Palo Alto, CA, USA). Samples were injected inthe split/splitless injector into a 30 m × 0.25 mm × 0.25 mm HP5-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 ata constant flux of 1 mL min−1. PAH quantification was conductedusing selected ion monitoring mode with the molecular ion of eachcompound at a minimum of 1.5 scan s−1.

PAHs were quantified relative to the perdeuterated PAHs intro-duced 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 con-centrations of, respectively, 210, 110, 210, 40 and 40 mg mL−1 inacetonitrile) were purchased from Sigma–Aldrich (France). Calibra-tion curves were established using a mixture of 16 parent PAHspurchased from LGC Standards (France), each compound at theconcentration of 100 mg mL−1.

For determining PAHs concentration in sea water, water sam-ples (1 L) were collected into Duran glass bottles that had beenpreviously heated at 500 ◦C to eliminate impurities. Samples wereextracted using Pestipur grade dichloromethane (3 × 100 mL perseawater sample). The combined organic extracts were dried byfiltering through anhydrous sodium sulphate (Na2SO4 Pestipurgrade) and concentrated to 2 mL by means of a Turbo Vap500 concentrator (Zyman, Hopkinton, MA, USA). Aromatic com-pounds were analyzed using the same GC–MS protocol previouslydescribed. The detection limit of this method was 1 ng L−1.

2.3.2. Water and tissue analyses of styreneStyrene concentration in water was directly monitored by UV-

spectrofluorometer (SF-UV) quantified by headspace (HS)/GC–MSusing the gerstel static HS module mounted on automated sampler.The combination of these two analytical techniques allowed to fol-low the styrene concentration in a qualitative way throughout theexposure and to evaluate the exposure concentration. Ten milliliterwater samples were stirred for 10 min at 70 ◦C. Styrene was thentransferred to a HP-5MS column (Hewlett-Packard, Palo Alto, CA,USA) using a 2.5 mL syringe. The split/splitless injector was usedin split mode at 250 ◦C. The oven program of temperature was thefollowing: from 40 ◦C (2 min) to 100 ◦C at 7 ◦C min−1, and, then to150 ◦C (2 min) at 25 ◦C min−1. The mass spectrometer was operatedin SIM mode as previously. Styrene was quantified with respect toethyl benzene using a calibration curve (from 0 to 10 mg L−1).

For tissues analysis, 1 mL of methanolic solution of ethyl ben-zene was added to 3 g of muscle per each fish. Then, sample wasplaced in the flask of the HS module and the same GC–MS protocolwas used.

2.4. Sequencing of target sequences and quantitative real-timeRT-PCR

For sequencing studies, juvenile turbots were collected in a farmin San Sebastian, Basque Country, Spain. Livers were frozen in RNAlater® and stored at −80 ◦C until processing. Fifty to a hundredmg of liver were homogenized in TRIzol® (Invitrogen, Carlsbad,

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 wasisolated for each individual fish liver following the manufacturer’sinstructions. Then, total RNA was used as template for cDNA syn-thesis by Super ScriptTM II reverse transcriptase PCR (Invitrogen,Leek, Netherlands) using random hexamers as primers and follow-ing the manufacturer’s recommendations in a conventional iCyclertermocycler (Bio-Rad, California, USA). cDNAs were PCR amplifiedusing degenerate primers, Fw-5′-CCA CAC TGA GAC CAG CAG-3′ andRv-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 CTGCAC (GC)AG C-3′ for ppar˛, designed through ClustalW alignmentsof known teleost and other phyla sequences to amplify conservedregions of ahr and ppar˛. These sequences were retrieved from theNational Center for Biotechnology Information (NCBI, U.S. NationalLibrary of Medicine, USA).

PCR was done with a 2 min activation and denaturing step at94 ◦C, followed by 35 cycles of a 30 s denaturing step at 94 ◦C and a30 s extension step at 72 ◦C. The annealing temperature was 57 ◦Cfor ahr and 55 ◦C for ppar˛.

PCR products were visualized in a 1.5% agarose gel, stainedwith ethidium bromide and purified using a PCR purification kit(Qiagen, Hilden, Germany). Single bands of ppar were directlysequenced, but whenever more than one band was amplified forahr, purified products were cloned using the TOPO-TA cloning kit(Invitrogen, Carlsbad, California, USA). PCR products for each genewere sequenced using degenerate primers while cloned insertswere sequenced using universal M13 primers in the Sequencingand Genotyping Service (SGIker) of the University of the BasqueCountry. Alignments and similarity matrices were performed usingBlastn, Blastx and ClustalW.

Transcription levels of cyp1a1 (GenBank ID: AJ310694), ahr andppar were measured in turbot livers by real time quantitative PCRusing TaqMan probes and primers. About 50–100 mg of fish liverwas homogenized in TRIzol®. Total RNA was isolated and cDNAwas obtained from 1 �g of total RNA by Super ScriptTM II reversetranscriptase PCR (Invitrogen, Leek, Netherlands) using randomhexamers as primers and following the manufacturer’s recommen-dations in the iCycler termocycler.

The real time PCR was run in 25 �L reactions on a 7003PCR machine (Applied Biosystems, California, USA) using TaqManReverse Transcription Reagent (New Jersey, USA). TaqMan probesand primers from turbot specific sequences were designed usingPrimer Express 3.0 software (Applied Biosystems, California, USA).All details are presented in Table 1. Universal conditions were usedin PCR for all genes: 1 cycle at 50 ◦C for 2 min, 1 cycle at 95 ◦C for10 min, 40 cycles at 95 ◦C for 15 s and at 60 ◦C for 1 min.

Amplified fragments were visualized after PCR in ethidiumbromide stained 1.5% agarose gels. Then, bands were cloned byTOPO-TA cloning kit and sequences were confirmed by sequenc-ing. Elongation factor 1 alpha (EF1-˛, GenBank ID: AF467776) wasused for normalization of transcription levels of target genes. Rela-tive transcript expression of a gene was calculated with the 2−��ct

method (Livak and Schmittgen, 2001) relative to the mean of con-trol animals sampled at day 3. Similar results were obtained whenthe mean value of the control group for each sampling day was used,but this approach did not allow studying the time-related variationin the transcription levels of control animals.

2.5. Acyl-CoA oxidase (AOX) activity

Livers were homogenized in 4 mL of TVBE buffer (1 mM sodiumbicarbonate, 1 mM EDTA, 0.1% ethanol and 0.01% Triton X-100, pH7.6) per gram of tissue, using a glass-Teflon homogenizer held inan ice bath. Homogenates were centrifuged at 500 g at 4 ◦C for15 min, and obtained supernatants were diluted 1:10 in TVBE buffer



Table 15′–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 dyeused for TaqMan real time PCR of studied genes.

Gene (GenBank accession no) Fw Rv Probe Product size (bp)

cyp1a1 (AJ310694) CTGCAAAGAGGGAGAGTATCTGAT GTTAGCCACAGACACAACAATGTAG FAM-CAGCTTCGACCCCTTCC-NFQ 100ahr (JN253594) TGGCGATGGCGAAGAGA TGTGTGACGACGAGGATGGT FAM-CGAGCTGATGTGGCG-NFQ 53ppar˛ (JN253593) AGAACATTGTGCAGGTGCTACAG ACGTGTCGTCCGGATGGT FAM-TCCACCTGCTGGCC-NFQ 56EF-˛ (AF467776) CCCCGGCGACAACGT CGCGGCGGATCTCCTT FAM-AACATCAAGAACGTGTCCGT-NFQ 59

and assayed for AOX activity. AOX activity was determined spec-trophotometrically measuring the H2O2-dependent oxidation ofdichlorofluorescein diacetate catalyzed by an exogenous peroxi-dase using palmitoyl-CoA 30 �M as substrate (Small et al., 1985).Measurements (� = 502 nm) were performed for four min duringthe linear phase of the reaction. Each sample was measured atleast twice and values accepted if the difference between them wasless than 10%. Total protein concentration was measured using theDC Protein Assay of Bio-Rad, based on the method of Lowry et al.(1951), with gamma-globulin as standard. AOX activity is given asmU AOX mg−1 protein.

2.6. Lysosomal membrane stability

At least 12 serial tissue sections (10 �m thick) of a small pieceof the liver obtained from fish were cut in a Leica CM 3000 cryostat(Leica Instruments, Nusslock, Germany) at a cabinet temperatureof −24 ◦C and then stored (not longer than 24 h) at −40 ◦C untilstaining.

Prior to staining, slides were air-dried at 4 ◦C for 20 min andanother 10 min at room temperature. Then, sections were intro-duced 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 minin a shaking water bath at 37 ◦C to destabilize the lysosomal mem-brane. Afterwards, sections were incubated for 10 min at 37 ◦C in0.1 M citrate buffer (pH 4.5) containing 2.5% NaCl, 0.04% naphtholAS-BI N-acetyl-�-d-phosphate dissolved in 2% dimethyl sulfoxideand 7% of POLIPEP® as a tissue stabilizer. After incubation, sec-tions were rinsed in 3% NaCl at 37 ◦C for 5 min in a shaking waterbath. Then, sections were transferred to 0.1 M phosphate buffer (pH7.4) containing 0.1% of diazonium dye Fast Violet B salt for 10 minat room temperature. Slides were rinsed in running tap water for10 min, fixed in 10% formaldehyde containing 2% calcium acetatefor 15 min at 4 ◦C and rinsed in distilled water. Finally, slides weremounted in Kaiser’s glycerin gelatin.

Labilization period (LP) was determined under a Leitz LaborluxS light microscope (40× objective) as the maximum accumulationof reaction product associated with lysosomes. Four determina-tions were made for each animal by dividing each section into fourapproximately equal areas and assessing the first peak of labiliza-tion period in each area (Zorita et al., 2008). A mean value was thenobtained for each section, corresponding to an individual liver.

2.7. Plasma vitellogenin (Vtg) levels

Blood samples were obtained using heparinized syringes. Aftera brief centrifugation (500 × g for 5 min), plasma supernatantwas quickly frozen in liquid nitrogen and stored at −80 ◦C untilfurther analysis. Plasma Vtg levels were measured by an indi-rect ELISA assay using a commercial polyclonal antibody againstturbot Vtg (CS-2, Biosense Lab., Norway), according to the pro-tocol described by Nilsen et al. (1998). Plasma samples werethawed on ice and diluted (1:50) in coating buffer (50 mMcarbonate-bicarbonate buffer, pH 9.6). Diluted samples (100 �L)were added in triplicate to 96-well microtiter plates and incubatedovernight at 4 ◦C. Plates were rinsed three times with phosphatebuffered saline solution containing 0.05% Tween 20 (PBST). Then,

wells were incubated for 1 h at room temperature with a blockingsolution composed of PBST supplemented with 1% bovine serumalbumin. After rinsing the plates with PBST, primary antibodydiluted 1:1000 in blocking solution was added (100 �L) to eachwell and incubated for 2 h at room temperature. Plates were thenrinsed with PBST and the secondary antibody (peroxidase conju-gated goat anti-rabbit IgG, Sigma Chemical Co., St. Louis, MO, USA)diluted in PBST (1:10,000) was added (100 �L) to each well andincubated for 1 h at room temperature. After rinsing the plates withPBST, visualization of detected Vtg molecules was done by addi-tion of 0.05 M phosphate–citrate buffer, pH 5 (100 �L) containing0.012% hydrogen peroxide and 0.4 mg mL−1 o-phenylenediamine.Incubation was performed for 30 min at room temperature in thedarkness. Reaction was stopped by adding 50 �L of 2 M H2SO4 toeach well. Absorbance was read at 492 nm. Non-specific bindingwas also measured for each plate, replacing samples by coatingbuffer alone. These wells were considered as blanks and averageabsorbance value of blanks from each plate was subtracted to eachsample well.

2.8. Gonad and liver histopathology

A portion of gonad tissue was fixed in 10% neutral buffered for-malin and routinely processed for paraffin embedding in a LeicaTissue processor ASP 3000 (Leica Instruments, Nussloch, Germany).Sections of 5 �m thickness were cut in a Leitz 1512 micro-tome (Ernst Leitz, Vienna, Austria), stained with hematoxylin/eosin(Wilson and Gamble, 2002) and examined under a Leitz LaborluxS light microscope (Wetzlar, Germany) for histopathological alter-ations such as gamete abnormalities. Gametogenic developmentalstages were determined for each animal, following the classifica-tion of Deng et al. (2007).

A piece of the liver from each sampled fish was processedas above and prevalence of histopathological alterations, such asmelanomacrophage centers (MMC), necrotic areas, vacuolizationor parasites was determined. The samples were also examined todetermine preneoplastic foci and neoplastic alterations accordingto the criteria established by Feist et al. (2004).

2.9. Statistical analyses

Statistical analyses were carried out with the aid of the SPSSstatistical package (V 14.0, SPSS Inc., Chicago, Illinois). For Vtglevels and AOX activity, differences along the exposure timewere studied by one-way analysis of variance (ANOVA) followedby the Duncan’s test for multiple comparisons between pairsof means. Significant differences between control and exposedgroups at each sampling time were studied using the Student’st test. Previous to the analysis, data were tested for normality(Kolmogorov–Smirnov normality test) and homogeneity of vari-ances (Levene’s test). In the case of LP and gene transcription levels,the non-parametric Kruskal–Wallis test was applied followed bythe Mann–Whitney U test. For histopathological data, the Chi-square test was used. In all cases, significance was established atp < 0.05.



Table 2Concentration of 16 EPA PAHs in samples of heavy fuel oil no. 6. Results are givenin �g g−1.

Compounds Mean ± SD

Naphthalene 686 ± 89Acenaphthylene 51 ± 3Acenaphthene 272 ± 14Fluorene 396 ± 20Phenanthrene 1936 ± 116Anthracene 213 ± 28Fluoranthene 125 ± 14Pyrene 516 ± 57Benzo[a]anthracene 213 ± 13Chrysene 464 ± 14Benzo[b + k]fluoranthene 81 ± 11Benzo[a]pyrene 167 ± 7Indeno[1,2,3-c,d]pyrene 16 ± 3Dibenz[a,h]anthracene 27 ± 5Benzo[g,h,i]perylene 48 ± 3

�PAHs 5211

3. Results

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

The results of the PAH analyses in the heavy fuel oil no. 6 sam-ple are shown in Table 2. Concentration ranged from 16 �g g−1 ofindeno[1,2,3-c,d]pyrene to 1936 �g g−1 of phenanthrene. The sumof the 16 PAHs reached 5211 �g g−1.

The PAH concentration in seawater along the experiment rangedfrom 311.43 to 4386.90 ng L−1 (Table 3). In the exposure tank,the highest PAH concentration was measured before fish wereintroduced in the tank (4386.90 ng L−1). However, the PAH con-centration 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 styreneconcentration in seawater fluctuated along the exposure periodreaching up to 0.7 mg L−1 (Fig. 2). Styrene concentration in thetissues of exposed fish was about 18.75 ± 12.03 �g g−1 whereasstyrene concentration in control fish was close to the GC–MS detec-tion limit.

0 70

0.80

0 40

0.50

0.60

0.70

ntr

atio

n (

mg

.L-1

)

0 10

0.20

0.30

0.40

Sty

rene

conce

n

0.00

0.10

0 20 40 60 80 10 0 12 0 140

Time (h)

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

3.3. Gene transcription levels

A fragment of 342 bp encoding ahr was sequenced in S. max-imus and predicted AHR protein sequence showed 86% aminoacid identity with AHR of Sebastiscus marmoratus (GenBank ID:ACH78368) (E value: 1e−51). For ppar˛, 200 bp were amplified fromturbot liver. The predicted protein sequence showed 95% aminoacid 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-˛, wereused to measure transcription levels of those genes by quantitativereal-time RT-PCR.

cyp1a1 transcription level was significantly increased in liverafter 3 and 14 days of exposure to heavy fuel oil. After the recoveryperiod, transcription levels returned to control levels (Fig. 3a). ahrtranscription level did not show significant variations between oil-treated and control turbots along the experiment, however highertranscription levels were measured in treated turbots than in con-trol animals after 14 days of exposure and 14 days in clean water(Fig. 3b).

Transcription level of ppar did not vary significantly betweenoil-treated and control animals along the experiment, but after therecovery period, transcription was significantly elevated in turbotspreviously exposed to heavy fuel oil in comparison to the controlgroup (Fig. 3c).

Table 3Concentration 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 52289.91 3499.63 1885.35 99.11 71.53 26.90 9.81Acenaphthylene 26.46 1.47 48.27 17.78 55.97 62.40 40.00Acenaphthene 4258.46 225.79 194.61 215.27 98.68 31.42 78.38Fluorene 4463.70 203.76 138.49 34.57 50.68 53.43 54.94Phenanthrene 708.43 356.66 142.27 30.36 13.89 17.65 32.55Anthracene 62.61 56.21 57.22 35.82 28.39 20.25 40.20Fluoranthene 6.86 5.57 0.00 8.81 3.94 19.10 5.44Pyrene 36.80 18.15 16.40 29.73 16.81 20.80 36.94Benzo[a]anthracene 229.05 9.06 121.78 4.18 2.16 2.57 4.90Chrysene 258.79 10.60 38.90 8.35 6.21 6.36 8.27Benzo[b + k]fluoranthene bdl bdl bdl bdl bdl bdl bdlBenzo[a]pyrene bdl bdl bdl bdl bdl bdl bdlIndeno[1,2,3-c,d]pyrene bdl bdl bdl bdl bdl bdl bdlDibenz[a,h]anthracene bdl bdl bdl bdl bdl bdl bdlBenzo[g,h,i]perylene bdl bdl bdl bdl bdl bdl bdl

�PAHs 62341.06 4386.90 2643.29 483.98 348.26 260.86 311.43



3

2

1a1pyc1a1pyc

30

40

Q

*

*

da

Fuel oil Styrene

0

1

3 da ys 7 da ys Recover y (7 da ys)

RQ

0

10

20

3 da ys 14 da ys Recovery (14 da ys)

RQ *

30

15

20

25

RQ

4

6

RQ

rharhab e

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10


0

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c f30

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10

15

RQ

*

c f

RQ

5

10

15

20

25

30

*

0


0


Control

Heavy fuel oil

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 aregiven 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 tran-scription of cyp1a1 and ahr genes (Fig. 3d and e), but up-regulatedsignificantly ppar transcription after 3 days of exposure (Fig. 3f).

3.4. Acyl-CoA oxidase activity

Compared to the control group, liver peroxisomal AOX activityincreased significantly after 14 days of exposure to fuel oil. Afterthe recovery period, AOX activity decreased significantly comparedto the control group and to the oil-exposed group at 14 days ofexposure (Fig. 4a). In fish exposed to styrene, AOX activity was sig-nificantly inhibited compared to controls only in the group sampledafter the recovery period (Fig. 4b).

3.5. Lysosomal membrane stability

In general, LP values were low in all fish from both heavy fuel oiland styrene experiments. Treated fish presented significantly lowerLP values than in the control group after 14 days of heavy fuel oilexposure (Fig. 5a) and 7 days of styrene exposure (Fig. 5b). Afterthe recovery period, LP values in both treated groups of turbotsincreased again up to control values.

3.6. Plasma vitellogenin levels

Levels of Vtg in the plasma of experimental fish were underdetection limit in most cases (data not shown). In the heavy fuel



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 indi-cate significant differences with respect to controls of the same day according to theStudent’s t test at p < 0.05.

oil experiment, only some animals in the control group at T0showed detectable Vtg levels in plasma. In the styrene experiment,some few females distributed in the different experimental groupsshowed detectable Vtg levels.

3.7. Gonad and liver histopathology

Both control and treated female turbots showed gonad at stage1 of gametogenesis (gonad composed of primary and secondaryoocytes) during the whole experiment, as corresponds to juvenileanimals. In male turbots, stages 1 (only spermatogonia) and 2 (sper-matogonia and spermatocytes) were observed. No differences ingamete development were detected between control and oil- or

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

ble

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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 3days. (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) Extensivenecrosis (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 foundin the gonad tissue of experimental animals.

The prevalence of the histopathological alterations found inthe liver of turbots is shown in Table 4. Livers with accumula-tion of MMC and focal necrosis were common in most groups,but prevalence of MMC was significantly lower in turbots exposedto oil for 14 days than in control fish sampled at the same time.Also in this sampling, the prevalence of extensive cell vacuoliza-tion (Fig. 6a) increased significantly in oil-exposed fish comparedto control fish and no recovery was observed (Table 4). Parasites(Fig. 6b) were common in both experimental groups, showing nodifferences between them. In the styrene experiment, no significantdifferences were found between treated and control animals forany of the pathologies studied, although for some of them, such asMMC (Fig. 6c) and necrotic areas (Fig. 6d), styrene-exposed animalsshowed higher prevalences in most samplings.

4. Discussion

This study aimed to determine the short-term effects of heavyfuel oil no. 6 and styrene, using a battery of molecular, cell andtissue-level biomarkers in juvenile turbots. The possible recoveryof biomarker responses was also evaluated by placing fish in cleanwater after exposure.

The concentration of measured PAHs in the samples of theheavy fuel oil no. 6 was 5211 �g g−1. This value is compa-rable 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 ofthe 16 EPA PAHs was 1782 �g g−1(Aas et al., 2000). The resultingPAH concentrations in the water column measured in this study(311.43–4386.9 ng L−1) were similar to those described after dif-ferent oil spills. For instance, 1 month after the Prestige oil spill,González et al. (2006) reported seawater PAH concentrations up to2090 ng L−1. Nevertheless, a wide range of PAH concentrations inseawater are described in the literature after oil spills: PAH con-centration reached values from 20.9 to 6000 ng L−1 after the Erikaand Exxon Valdez oil spills, respectively (Short and Harris, 1996;Tronczynski et al., 2004).

At the beginning of the experiment, before fish were introducedin the tank, the PAH concentration in seawater was 4386.9 ng L−1.After two days of exposure, PAH concentration in seawaterdecreased approximately ten times (483.98 ng L−1). This decline ofPAH concentration is explained by the volatilization of low molecu-lar weight PAHs, such as naphthalene, as well as by the introductionof fish that are able to take up PAHs. In a similar study where mulletswere exposed to fresh heavy fuel oil, the PAH concentration in sea-water 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 ofup to 0.7 mg L−1. Information on styrene levels in the aquatic envi-ronment is limited. Styrene concentrations in water are usually lessthan 20 �g L−1 (Alexander, 1997). Nevertheless, concentrations up



to 100 �g L−1 have been detected in a small number of industrialeffluents in the US (Alexander, 1990). Information on the toxic-ity of styrene to aquatic organisms is restricted to acute studies,as styrene is a highly volatile compound and its half-life in wateris 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 at96 h, respectively, indicating a moderate toxicity for these ani-mals (Cushman et al., 1997). Mamaca et al. (2005) showed thatstyrene concentration in seawater decreased from the nominalconcentration of 2–0.2 mg L−1 during Symphodus melops expo-sure. Volatilization and biotransformation are the main processesaccounting for the loss of styrene when it reaches the environ-ment (Fu and Alexander, 1992; Gibbs and Mulligan, 1997). In spiteof the low styrene concentration in seawater the results of thisstudy 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 detectedconcentration was high (18.75 �g g−1) when compared with a pre-vious study in which styrene concentration in crab tissues was0.093 �g g−1 7 days after the Ievoli Sun spill (Law et al., 2003).

Biotransformation is among the most studied processes inaquatic animals exposed to organic xenobiotics. Transcriptionalinduction of cyp1a1 in fish may serve as a sensitive marker of xeno-biotic exposure and early biological response (Hahn and Stegeman,1994). Induction of cyp1a1 transcription has been observed infish following laboratory exposure to PAH compounds, sedimentextracts, and also in fish captured from PAH-contaminated envi-ronments (Bilbao et al., 2010; Courtenay et al., 1999; Meucci andArukwe, 2006; Paetzold et al., 2009). In the present study, cyp1a1transcription was significantly induced in liver after 3 and 14 daysof exposure to heavy fuel oil. This response has also been observedin mullets after exposure to Prestige oil (Bilbao et al., 2010). After14 days in clean water, cyp1a1 transcription level returned to con-trol levels, which indicates the reversibility of the response afterinducing compounds have been removed from water.

In mammals, styrene degradation is also catalyzed bycytochrome P450-dependent monoxygenases (Gibbs and Mulligan,1997; Rueff et al., 2009). However, in this work no changes wereobserved in cyp1a1 liver transcription levels after exposure tostyrene.

It is well documented that cyp1a1 transcription in responseto 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 (ahr1and ahr2) (Andreasen et al., 2002; Hahn et al., 1997; Hansson et al.,2003). Although the respective functions of these two ahr forms arenot well understood, ahr2 appears to play a major role in mediatingthe toxicity of PAHs (Clark et al., 2010). The ahr sequence obtainedin turbot was more similar to ahr1 than ahr2 from different fishspecies. In our study, heavy fuel oil and styrene exposure did notproduce any significant change in ahr transcription levels at themeasured timepoints, although exposed turbots presented higherlevels than control animals.

PAHs have been described to produce peroxisome proliferationin several aquatic organisms (Cajaraville et al., 2003; Oakes et al.,2005; Ortiz-Zarragoitia and Cajaraville, 2005; Ortiz-Zarragoitiaet al., 2006). Similar to mammals, peroxisome proliferators act infish by binding to PPARs (Colliar et al., 2011). PPARs are membersof the nuclear receptor superfamily of ligand-activated transcrip-tion factors (Mandard et al., 2004). Target genes, like those ofthe peroxisomal �-oxidation pathway in the case of ppar˛, aremainly involved in lipid homeostasis (Mandard et al., 2004). Tur-bots exposed to heavy fuel oil showed higher levels of hepatictranscription of ppar� compared with controls, but this differ-ence was statistically significant only after 14 days in clean water.This could indicate that previous oil exposure can affect ppar˛

transcription and that effects could be observed even when thecontaminant was removed. In case of styrene, ppar transcriptionlevel was significantly up-regulated after 3 days of exposure, sug-gesting that the response might appear more rapidly after exposureto styrene than to heavy fuel oil.

One of the genes whose transcription is regulated by ppar isAOX (Dreyer et al., 1992), the first and rate limiting enzyme of theperoxisomal fatty acid �-oxidation pathway. AOX is generally usedas a marker of peroxisome proliferation since its activity is inducedwhen sensitive organisms are exposed to peroxisome prolifera-tors (Cajaraville et al., 2003). Thus, induction of AOX activity hasbeen 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 after14 days of exposure to heavy fuel oil in comparison with controlvalues. In grey mullets (Chelon labrosus) exposed to Prestige-likefuel oil, AOX activity was induced after 2 and 16 days of expo-sure (Bilbao et al., 2010). In contrast, we did not find differencesin AOX activity in animals exposed for up to 7 days to styrene,suggesting that styrene did not act as a peroxisome proliferatorin turbot liver under the assayed conditions. Thus, the significantup-regulation of ppar transcription after 3 days of styrene expo-sure, was not followed by an induction in AOX activity. Similarly, inmullets, up-regulation of ppar transcription after 2 days of expo-sure to weathered fuel oil was not immediately reflected in inducedAOX activity. Increased AOX activity was not recorded until day 16,suggesting that observed regulation in transcription levels of theppar receptor needed more time to elicit a response at proteinactivity level (Bilbao et al., 2010).

Lysosomes have a crucial role in the detoxification of toxic sub-stances and constitute an important target of toxicants effects. Forthese reasons, changes in the lysosomal system have been usedas general marker of pollutant impact in a number of field studiesusing fish as sentinel organisms (Köhler et al., 2002; Zorita et al.,2008). Both control and exposed turbot of this study presentedLP values below 10 min, which indicate a bad condition of fishaccording to established baseline and critical values for environ-mental health assessment (Köhler et al., 2002). Nevertheless, lowLP values (6–7 min) have been reported in eelpout and in floun-der collected from reference sites in the Baltic Sea (Barsiene et al.,2006; Lehtonen et al., 2006). The range of LP values reported in con-trol fish could be attributed to species differences or to differencesin sample preparation between laboratories. Even so, lysosomalmembrane stability was significantly reduced in hepatocytes of oil-and styrene-exposed turbots in comparison with control turbotsafter 14 days and 7 days of exposure, respectively. This indicatesthat tested contaminants were taken up and produced deleteriouseffects in treated fish. After the recovery period in clean water, LPvalues of oil- and styrene-exposed turbots increased up to controlvalues, suggesting that the depuration period in both experimentswas enough to enable them to recover the lysosomal membranestability up to control values. In agreement with our results, thicklipgrey mullets exposed under laboratory conditions to Prestige-likefresh and weathered heavy fuel oil for 2 and 16 days showed signif-icantly lower LP values than the controls (Bilbao et al., 2010). Thus,lysosomal membrane stability has to be considered as an indica-tor of non-specific general physiological stress and may be used toassess the recovery of health impairment produced by fuel oil andstyrene.

Vtg levels have been widely used as a biomarker of exposureto xenoestrogens in fish (Arukwe and Goksøyr, 2003). Levels ofVtg in plasma of experimental fish were under detection limitin most cases, suggesting that heavy fuel oil and styrene did notact as typical xenoestrogens in turbots. Zebrafish exposed to thewater accommodated fraction of crude oil showed antiestrogenic



effects (Arukwe et al., 2008). The antiestrogenicity of PAH (3-methylcholanthrene) has also been reported in rainbow trouthepatocytes in vitro (Navas and Segner, 2000). Similarly, immaturerainbow trout exposed to naphthalene showed reduced plasma17�-estradiol levels and reduced reserves directed to vitellogene-sis and gamete development (Tintos et al., 2006). In vitro ER bindingassays showed a low affinity of styrene for the estrogen recep-tor. Besides, more specific in vitro reporter gene assays and in vivorat uterotrophic assay did not show estrogenic effects caused bystyrene oligomer exposure (Ohno et al., 2003). Accordingly, con-trol and exposed turbots in both experiments showed gametes atearly stages of development (stage 1) during the whole experiment.No differences in gamete development existed between controland exposed fish, indicating that reproductive potential was notdisturbed by heavy fuel oil or styrene exposure.

Strong evidence for a relationship between water pollution andfish disease has been found by several researchers (Grinwis et al.,2000; Lang et al., 2006; Stehr et al., 1998). The occurrence of liverlesions considered to be associated with exposure to anthropogeniccontaminants has been recorded in fish from several coastal areas(Lang et al., 2006; Myers et al., 1998; Stentiford et al., 2003; Vethaakand Jol, 1996). Laboratory experiments with PAHs performed withseveral fish species have also resulted in different histopatho-logical alterations (Bilbao et al., 2010; Reynolds et al., 2003). Inour study, control animals of both experiments presented sev-eral histopathological alterations including accumulation of MMC,necrosis, vacuolization and presence of parasites. These alterationswere also observed at the beginning of the experiments, whichcould be related to the low LP values observed. A significant increaseof extensive cell vacuolization was observed at 14 days and afterrecovery period in the liver of turbots exposed to fuel oil. Thisfinding is similar to those reported for flounders caught from pol-luted UK estuaries (Lyon et al., 2004) and for salmons after 10days of exposure to the water soluble fraction of Alaska NorthSlope crude oil (Brand et al., 2001). Therefore, it could be suggestedthat fuel oil exposure caused effects in fish liver at short time thatremained after 14 days in clean water. Prevalence of MMC in turbotsdecreased significantly after 14 days of fuel oil exposure, in agree-ment with previous studies on fuel oil effects in mullets (Bilbaoet al., 2010). In flounders, chronic exposure to PAH also caused a sig-nificant decreased in MMC amount (Payne and Fancey, 1989) andreduced activity of melanomacrophage aggregates was observedin individuals from polluted areas (Broeg, 2003), indicating severeeffects of PAHs on the immune system of chronically exposed fish.No significant liver alterations were observed in turbots exposedto styrene although a trend for a higher prevalence of MMC, areasof necrosis and focal cell vacuolization was identified in styrene-exposed turbots compared to control animals. Recovery was onlyobserved in focal cell vacuolization after 7 days in clean water.

As a summary, transcript of the PAH-responsive gene cyp1a1was significantly overexpressed after exposure to heavy fuel oil.Peroxisomal AOX activity was induced 14 days after exposure. Bothresponses disappeared after withdrawal of the contaminant. Lyso-somal membrane stability test gave consistent results with reducedLP after 14 days of exposure to heavy fuel oil and further recovery.As indicated by Vtg data, heavy fuel oil exposure did not producexenoestrogenic effects in turbots. Histopathological examinationof gonad indicated that heavy fuel oil did not provoke effects ongamete development. However, significant histopathological alter-ations were observed in liver of oiled turbots. Styrene did not inducealterations in transcription levels of genes related to biotransfor-mation whereas ppar transcription level was up-regulated after3 days exposure. However, styrene did not behave as a typicalAOX inducing peroxisome proliferator in turbot. Styrene provokeda general stress: however, we did not observe evidences of effectson gamete development and liver histology. Finally, according to

the results of the present study, the suite of selected biomarkersproved to be useful to determine the early impact of, and recoveryfrom, exposure to heavy fuel oil and styrene in turbot.

Acknowledgements

This work was supported by the European Commission(Directorate-General Environment) through the PRAGMA project“A pragmatic and integrated approach for the evaluation of envi-ronmental impact of oil and chemicals spilled at sea: Input toEuropean guidelines” (grant no. 07.030900/2005/429172/SUB/A5),the Spanish MEC (project CANCERMAR, CTM2006-06192 and a pre-doctoral fellowship to P. Ruiz) and the Basque Government throughthe strategic action ETORTEK-IMPRES and a grant to consolidatedresearch groups (GIC07/26-IT-393-07). Work funded by the Uni-versity of the Basque Country (UFI 11/37).

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