Biochemical response in gilthead seabream (Sparus aurata) to in vivo exposure to a mix of selected PAHs

ARTICLE IN PRESS Ecotoxicology and Environmental Safety 72 (2009) 1296–1302

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Biochemical response in gilthead seabream (Sparus aurata) to in vivo exposure to a mix of selected PAHs Justyna Kopecka-Pilarczyk a,, Ana Dulce Correia a,b a b

CIIMAR-Centre for Marine Environmental Research, Laboratory of Environmental Toxicology, Rua dos Bragas, 289, 4050-123 Porto, Portugal Unidade de Bioquı´mica Fı´sica, Instituto de Medicina Molecular, Faculdade de Medicina de Lisboa, Av. Prof. Egas Moniz, 1649-028 Lisboa, Portugal

a r t i c l e in f o

a b s t r a c t

Article history: Received 12 March 2008 Received in revised form 16 November 2008 Accepted 6 December 2008 Available online 23 January 2009

In our previous studies we assessed the biochemical response of juvenile seabream to the exposure to three selected polycyclic aromatic hydrocarbons (PAHs) (phenanthrene, pyrene and fluorene) in shortterm laboratory experiments. Since in the analyses conducted in the field it is much more natural to expect the presence of multiple contaminants at the same time, the objective of this study was to analyse the reaction of seabream to the exposure to a mix of the three PAHs tested previously. The proportions of the individual components in the PAH mix used in our experiments were constant and determined on the basis of the concentration addition concept. We measured a set of hepatic biomarkers: 7-ethoxyresorufin-O-deethylase (EROD), glutathioneS-transferase (GST), catalase (CAT), superoxide dismutase (SOD) and lipid peroxidation; as well as biliary fluorescent aromatic compounds (FACs) metabolites. Moreover, we calculated the biotransformation index, the condition factor and the hepatosomatic index. Only a few of these indicators provided statistically significant response to the tested exposures, while FACs showed very strong increase with increasing concentration of the tested mix of PAHs. The results of this research are expected to contribute to the establishment of a good biochemical index of exposure to mixtures of PAHs in laboratory experiments, which can be further useful in field studies dealing with the impact of organic pollutants on marine organisms. & 2008 Elsevier Inc. All rights reserved.

Keywords: Biomarkers Gross indices FAC PAH mix Seabream Short-term exposure Experiment in vivo

1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are a group of ubiquitous environmental pollutants containing two or more aromatic rings. Although natural sources of PAHs exist (e.g., forest fires, natural petroleum sources), the contamination of aquatic environments with these substances originates mainly from anthropogenic sources such as incomplete combustion of fossil fuels, petroleum spills, discharge from ships, industrial effluents, highway runoff from creosote-treated wood products and creosote spills (reviewed in Meador et al., 1995). PAHs have a strong capacity to bioconcentrate, but they generally do not bioaccumulate to their full potential because many of them can be metabolised. Nevertheless, these pollutants either individually or in combination may have sub-lethal effects at the cellular, organ, or individual level, e.g., causing changes in the enzymatic, genetic, behavioural and reproductive activities (Allen and Moore, 2004).

 Corresponding author. Fax: +351 22 339 06 08.

E-mail address: [email protected] (J. Kopecka-Pilarczyk). 0147-6513/$ - see front matter & 2008 Elsevier Inc. All rights reserved. doi:10.1016/j.ecoenv.2008.12.003

Many biologically active PAHs volatilise or are rapidly metabolised, so direct measurement of contaminant burdens may not yield an accurate estimation of contaminant exposure or uptake. Therefore, indicators of biological effects (biomarkers) have been introduced to supplement chemical analyses (McCarthy and Shugart, 1990; Peakall, 1992). Specifically, effects recognised at the biochemical level are generally used as early warning signals for assessing the influence of contaminants on organisms (Haux and Fo¨rlin, 1988). Biomarker responses to petroleum PAHs, dominated by twoand three-ringed aromatics, can be quite different from the response to pyrogenic PAHs, dominated by four- and five-ringed aromatics (Anderson and Lee, 2006). Therefore, in order to provide results most relevant from the point of view of ecotoxicology, in our research we focussed on a selection of those ‘‘less toxic’’ PAHs: phenanthrene (Phe), pyrene (Pyr) and fluorene (Fluo). Phe, Pyr and Fluo occur in fossil fuels and they are present in products of incomplete combustion. Fluo has 2 benzene rings, Phe has 3 benzene rings, while Pyr has 4 benzene rings, and all three are considered non-carcinogenic PAHs, in contrast to most high molecular weight PAHs. Nevertheless, their impact on the environment is so profound that they are included in the EPA’s priority pollutant list (US EPA, 1987).

ARTICLE IN PRESS J. Kopecka-Pilarczyk, A.D. Correia / Ecotoxicology and Environmental Safety 72 (2009) 1296–1302

PAHs often occur in the environment as complex mixtures composed of up to 100 different compounds, and each compound has its unique toxic potency (Santodonato et al., 1981). However, as it was pointed out by Anderson and Lee (2006), many studies concern either exposures to single compounds in the laboratory or to mixtures of various contaminants in the field, making it difficult to closely tie the measured responses exclusively to petroleum exposure. In order to make a step towards filling this gap, in the present study the biochemical response of a particular fish species was evaluated as a result of exposure to the mixture of the three PAHs (Phe, Pyr and Fluo) previously analysed individually in Correia et al. (2007) and Kopecka-Pilarczyk and Correia (submitted). The concentrations of individual compounds in the mixture was determined on the basis of a classical concept of concentration addition (CA) which is used in aquatic toxicology to predict joined toxic effects of ‘‘similarly acting’’ chemicals (Faust et al., 2001). In our case this concept was applied to determine proportions of the three PAHs for which their individual effect in the mixture adds up at the same level. The CA model was also used to determine the specific concentrations of the mix of PAHs in the experiments in such a way that the combined effect be comparable with the effects attained in our previous studies with the individual PAHs. Laboratory experiments were conducted with juvenile gilthead seabream (Sparus aurata), which is a euryhaline and eurythermal species widely cultured in Europe (Pretti et al., 2001). Laboratory studies are useful for evaluating the response of organisms to specific contaminants at known concentrations in controlled conditions. This results in better understanding of cellular impact of selected chemicals, which is necessary for choosing the most useful parameters to follow in field studies (Legeay et al., 2005). In comparison to the work of Correia et al. (2007) and KopeckaPilarczyk and Correia (submitted), where laboratory experiments with selected individual PAHs were conducted, in the present study we investigated the effect of a mixture of the PAHs analysed there. A suite of biomarkers reflecting exposure and effects of contaminants at different biological levels were selected for this study in order to provide a comprehensive overview of the impact of the chemicals on the fish. These biomarkers are described in the remainder of this section. Biotransformation reactions are most abundantly studied in the liver. The detoxification process of organic contaminants such as PAHs is typically divided into two phases: biotransformation (Phase I) and conjugation (Phase II). Phase I is represented by the activity of CYP1A, measured commonly as 7-ethoxyresorufinO-deethylase (EROD) activity. Phase II is reflected in the activity of glutathione-S-transferase (GST) (Stegeman et al., 1992). These two biomarkers were chosen for our research. Phase II enzymes can play an important role in homeostasis as well as in detoxification and clearance of many xenobiotic compounds. However, compared with Phase I systems, the induction responses of Phase II enzymes are generally less pronounced (reviewed by van der Oost et al., 2003). Additionally, the biotransformation index (BTI) defined as the ratio between Phase I (EROD) and Phase II (GST) activities provides valuable indication of the balance between bioactivation and detoxification (van der Oost et al., 1998). Defence systems that tend to inhibit formation of oxyradicals include the antioxidant enzymes such as superoxide dismutase (SOD) and catalase (CAT). It is known that SOD converts the superoxide anion (O 2 ) to hydrogen peroxide (H2O2), and H2O2 is further converted to water (H2O) and oxygen (O2) by CAT (Winston and Di Giulio, 1991). Antioxidants are generally less responsive to pollutants than Phase I or Phase II enzymes, but they also participate in detoxification processes (van der Oost et al.,

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2003) and play a crucial role in maintaining cell homeostasis. Lipid peroxidation (LP) is a very important consequence of oxidative stress and cellular damage, and refers to the oxidative deterioration of polyunsaturated lipids (Stegeman et al., 1992). Following metabolism by Phase I and Phase II detoxification enzymes, the major route of excretion of high molecular weight metabolites is the bile. Analyses of PAH metabolites in fish bile are a sensitive method for detection of PAH contamination from both pyrolytic and petroleum origin. Although the biliary metabolite content does not represent an adverse biological effect by itself, significant correlations have been established between the presence of hepatic lesions and mean concentrations of biliary metabolites in fish (Gagnon and Holdway, 2000). Some of the PAHs can be excreted directly as unconjugated polar metabolites, but most PAHs will be excreted into the bile after conjugation by Phase II enzymes (reviewed by van der Oost et al., 2003). Therefore, in addition to the biomarkers discussed earlier, PAH metabolites in bile measured as fluorescent aromatic compounds (FACs) were also selected in our study in order to observe the response of fish to the exposure to the mix of PAHs more closely. Moreover, in our study two gross morphometric indices supplementing the biomarker analyses were calculated: the condition factor (CF) which helps in assessing the general condition of fish, and the hepatosomatic index (HSI) which may be used to detect a possible liver disease. These indices can provide important information on potential pollution impact on fish (van der Oost et al., 2003).

2. Materials and methods 2.1. Test organisms and the design of experiments Juvenile seabream (ca. 5 months old) were obtained from TIMAR Lda. (Setu´bal, Portugal). Until the start of exposures, the fish were kept in 2200 l fibre aquaria supplied with filtered seawater (3572 ppt) and fed with food pellets (Aquasoja, Portugal). Fish used to conduct independent exposure studies with the mix of PAHs had an average body length (cm) and weight (g), respectively, in the following ranges: in the 1st study: 5.7470.25, 2.2470.31; in the 2nd study: 6.2370.26, 3.0070.32. Waterborne exposures were conducted in 17 l glass aquaria under semi-static conditions. The experiments were conducted at 1671 1C in filtered seawater (3572 ppt) under a photoperiod of 12 h light:12 h dark. Five randomly chosen fish were acclimatised to each test aquarium for 24 h with aeration (pre-exposure phase). Afterwards, fish were exposed to the test chemicals for 4 days. Always five aquaria per treatment and solvent control (acetone) were used in each experiment. Food was not provided during the acclimation and in the course of exposures. Master stock solutions of chemicals were prepared in acetone (analytical grade) and stored at 20 1C. The chemical solutions were regularly prepared from aliquots of the master solution and administrated directly into the aquaria. Exposures were daily renewed along with natural filtered seawater (50% of total volume), and the solvent concentration in the tank never exceeded 0.0014%. Dissolved oxygen saturation (480%) and total ammonia concentrations (o0.5 mg l1) were monitored every 2 days. The effects of the mix of PAHs on fish were assessed at two exposure concentrations in each study, with one concentration being retested in the second study. Thus pooled effect information for three concentrations was gathered. Two consecutive studies with a mix of PAHs were conducted with the following nominal concentrations of the PAH mix (control was run in parallel in each study):

 1st study: 0.148 mM (26 mg l1) and 0.444 mM (77 mg l1);  2nd study: 0.148 mM (26 mg l1) and 0.222 mM (39 mg l1). The mix of PAHs was prepared in the following proportions: 34% Phe, 10% Pyr, 56% Fluo, based on the CA model; the details are explained in Gonc- alves et al. (2008).

2.2. Chemicals Phe (X97% purity), Pyr (98% purity) and Fluo (98% purity) were purchased from Sigma-Aldrich (St. Louis, USA) and they were of analytical grade.

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2.3. Preparation of samples After 4 days of exposure fish were sacrificed, and liver (50–100 mg wet weight) and bile were collected (the latter by incising the gall bladder) and stored in cryovials at 80 1C until the analyses. For hepatic enzymes two pooled liver samples were processed, and for FACs individual bile samples were analysed. Additionally, the total body length (cm) and weight (g), as well as liver weight (g) of each individual fish were measured. The total number of fish analysed for each biomarker or morphometric index is given in Table 1.

2.4. Biochemical assays Livers were homogenised in ten volumes of phosphate buffer (100 mM, pH 7.5) containing 1 mM EDTA, and afterwards centrifuged at 10,000g for 20 min at 4 1C. The resulting postmitochondrial supernatants (S9 fraction) were distributed into aliquots and stored at 80 1C prior to analyses. Sample volume of 250 ml was used for the lipid peroxidation assay and 40 ml for the enzymatic measurements (EROD, GST, CAT and SOD).

2.4.1. Protein The total protein concentration was determined in the S9 fraction according to the method of Lowry et al. (1951) adapted to microplates, using BSA (bovine serum albumin) as a reference standard. Then samples for EROD were diluted to obtain 4 mg ml1 protein, and for SOD, GST and CAT—2 mg ml1 protein. In the samples diluted in this way, the biomarkers listed above were analysed.

2.4.2. EROD The procedure of Sole´ et al. (2000) using the S9 fraction was applied, and measured with a fluorimeter. The S9 fraction liver samples (25 ml) were incubated at 30 1C for 10 min in a final volume of 0.5 ml containing phosphate buffer (87 mM, pH 7.5), 0.22 mM NADPH and 3.70 mM 7-ethoxyresorufin. The reaction was stopped by adding 1 ml of ice-cold acetone, and the samples were centrifuged at 400g. Then 7-hydroethoxyresorufin fluorescence was determined at the 530/585 nm excitation/emission wavelengths. EROD activity was expressed as pmol min–1 mg prot–1.

2.4.3. GST In order to determine GST activity, the procedure of Habig et al. (1974) modified by Frasco and Guilhermino (2002) to microplates using 1-chloro-2,4dinitrobenzene (CDNB) as substrate was applied. Reaction mixtures contained 4.95 ml phosphate buffer (0.1 M, at pH 6.5):0.9 ml GSH (10 mM):0.15 ml CDNB (60 mM). In the microplate, 0.2 ml of the reaction mixture was added to 0.1 ml of the sample, with the final concentration of 1 mM GSH and 1 mM CDNB in the assay. The activity rate of GST was measured as the change in OD/min at 340 nm (ext. coefft. 9600 M1 cm1) and expressed as nmol min–1 mg prot–1.

2.4.5. SOD The protocol of McCord and Fridovich (1969) adapted to microplates by Ferreira et al. (2005) was used. SOD activity was determined in the S9 fraction as inhibition of the cytochrome c reduction at 550 nm. The reaction contained phosphate buffer (50 mM, pH 7.8), 50 mM hypoxanthine, 1.98 mU ml–1 xanthine oxidase and 10 mM cytochrome c. The relative activity in units of SOD were measured (U mg prot–1), with one unit SOD being the amount of sample causing 50% inhibition of the cytochrome c reduction under the standard conditions of the assay.

2.4.6. LP The protocol of Niki (2000) was used to determine LP (malondialdehyde-MDA equivalents) in the S9 fraction and measured on microplates. Subsamples of tissue homogenate were incubated with 100% TCA, and after centrifugation the supernatant was incubated for 30 min at 100 1C with 1% TBA, 0.05 M NaOH and 0.025% BHT. Then the absorbance of this solution was measured at 532 nm (e ¼ 1.54  105 M–1 cm–1). The lipid peroxidation was expressed as MDA equivalents per g liver (w.w.). The analyses of proteins, GST, SOD and LP were carried out using a microplate reader (BIO-TEK PowerWave 340), CAT—cuvette spectrophotometer (LKB BIOCHROM, Ultrospec II), and EROD—cuvette fluorimeter (BIOTEK instruments, SFM 25).

2.4.7. Phenanthrene-, pyrene- and fluorene-type metabolite analysis (FACs) Bile samples from control were diluted in 48% ethanol to 1:1500, and bile samples from exposed fish to further dilutions (1:5000, 1:10,000 and 1:100,000). Fluorescent readings were made at 260/380 nm (excitation/emission) for Phe-type metabolites, 341/383 nm for Pyr-type metabolites (Krahn et al., 1993) and 275/ 328 nm for Fluo-type metabolites. The pair excitation/emission for Fluo was optimised after having followed the general descriptions by Chetiyanukornkul et al. (2006). The congener PAHs were used as reference standards. A 5-nm slit width was used for excitation and emission. Measurements were performed on a BIOTEK SFM25 fluorimeter. To avoid fluctuations of the spectrofluorimeter and to allow samples to be comparable, a calibration curve was made for each metabolite, using the following standards: 1-hydroxypyrene (10, 5, 2.5, 0.625 and 0.156 mg l1), Phenanthrene (400, 100, 25 and 2.5 mg l1) and 1-hydroxyfluorene (100, 50, 10, 1 mg l1). Biliverdin content was estimated spectrophotometrically at 380 nm in the same samples and used to normalise FACs concentrations.

2.5. Gross morphometric indices and the BTI For each individual seabream fish gross morphometric indices were calculated according to the following formulas (van der Oost et al., 2003): condition factor ðCFÞ ¼ ðW T =L3 Þ100,

2.4.4. CAT The method described by Claiborne (1985) was used for the determination of CAT activity, and CAT was measured in cuvettes using a spectrophotometer by the decrease in absorbance at 240 nm because of H2O2 consumption (e ¼ 40 M1 cm1). The results were expressed as mmol min–1 mg prot–1.

hepatosomatic index ðHSIÞ ¼ ðW H =W T Þ100, where W T is the total wet weight (g), L the total length (cm) and WH the liver weight (g). The biotransformation index (BTI) was calculated as the ratio of EROD to GST (van der Oost et al., 1998).

Table 1 The level of biomarkers: EROD, GST, BTI, CAT, SOD [x  SD] and LP, FAC metabolites, gross indices: HSI, CF [x  SD] and the number of replicates (n) in juvenile seabream exposed to the mix of PAHs (Phe+Pyr+Fluo). Endpoint

First study

Second study 0.148 mM

Control 1

1

EROD (pmol min mg prot ) GST (nmol min1 mg prot1) BTI (pmol nmol1) CAT (mmol min1 mg prot1) SOD (U min1 mg prot1) LP (nmol MDA g1 w.w.) HSI (%) CF (%) Phenanthrene metabolites (mg ml1) Pyrene OH-metabolites (mg ml1) Fluorene OH-metabolites (mg ml1)

98.2677.80 149.4712.70 0.7170.09 40.1573.48 16.2870.97 34.91712.65 1.3270.19 1.2070.09 170.3738.81 22.59710.87 44.16710.63

(9) (10) (9) (10) (10) (10) (25) (25) (14) (14) (15)

63.3277.15* 118.677.38 0.5570.06*** 44.5670.77 14.8870.39 47.92712.67 1.2870.25 1.1470.07* 506171967** 10007362.5** 10567485.0**

0.444 mM (9) 49.3274.57*** (10) 160.3711.65 (9) 0.3270.03 (10) 45.1571.52 (10) 14.2970.29 (10) 48.44713.82 (25) 1.4070.31 (25) 1.1770.07 (13) 1805373638*** (14) 31947704.4*** (13) 517571823***

0.148 mM

Control (10) (10) (10) (10) (10) (10) (25) (25) (14) (14) (15)

41.7973.03 141.374.23 0.3070.07 36.5571.57 12.3270.33 47.8778.08 1.1870.22 1.2270.10 237.27100.2 25.387515.03 56.03710.71

(9) (10) (9) (9) (10) (8) (25) (25) (14) (13) (14)

50.3776.70 125.378.89 0.4170.17 36.8572.76 14.5670.61** 49.68711.08 1.1970.22 1.2270.10 402071162** 917.57277.3** 859.87314.3**

0.222 mM (10) (10) (10) (10) (10) (9) (25) (25) (14) (15) (15)

53.2577.40 130.877.78 0.3970.13 38.7670.98 13.6870.42 52.49712.16 1.2670.18 1.2870.12 1197274069*** 23677731.0*** 351171061***

Statistically significant differences are indicated with asterisks: * po0.05, **po0.01, ***po0.001. xFmean, SD—standard deviation, SDFSD of the mean.

(8) (10) (8) (9) (10) (10) (25) (25) (14) (14) (13)

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2.6. Statistical treatment of data Biomarkers (EROD, GST, CAT, SOD) were expressed as xðmeanÞ  SD (standard deviation of the mean), whereas LP, FACs, CF and HSI were expressed as xðmeanÞ  SD (standard deviation). The differences in biomarkers in comparison to control for each study were assessed by one-way ANOVA, the Kruskal–Wallis non-parametric rank test, followed by Dunn’s non-parametric post hoc multiple comparison test for selected pairs, designed for data sets with no30. In a more typical approach, the data would be log transformed to achieve normal distribution, verified using the Shapiro–Wilk test, with homogeneous variance, checked with Levene’s test. Unfortunately, a few tested cases failed this verification, so the parametric ANOVA tests could not be applied reliably to our data. Linear Pearson correlations were used to examine dependencies between all the enzymes. All statistical calculations were completed using the StatSoft Statisticas 6.0 software package, while the Dunn’s test calculations were conducted using GraphPad InStat.

2.7. Ethics statement Animals used in the experiments conducted for the purpose of the research described in this paper were treated in accordance with the Portuguese Animal Welfare Law passed in 1995 by the Parliament of Portugal.

3. Results The results of measured variables obtained for seabream exposed to the mix of PAHs (Phe, Pyr and Fluo) as well as the numbers of fish per group (n) are gathered in Table 1. Statistically significant differences for each variable with respect to control for each study found by Dunn’s non-parametric post hoc test are indicated with asterisks. Relative values of the variables, normalised with respect to control at time 0 which was set at level 0, are illustrated in Figs. 1 and 2. The results are presented in both tabular and graphical form in order to provide the reader with both access to the accurate data and suggestive visualisation of the pattern observed for each measured component. The biomarker results obtained for seabream exposed to the mix of PAHs are very variable (see Table 1 and Figs. 1 and 2) and only in a few cases are statistically significant, mainly for FAC metabolites (Fig. 2). The EROD and SOD activities as well as the BTI were well inhibited in the first run of the experiment. In the second run they were slightly induced, except for SOD for which considerable inhibition was noticed for the lowest PAH mix concentration. The GST activity slightly increased in comparison to control only for the highest tested concentration (0.444 mM), while in the other cases (0.148 and 0.222 mM) it was inhibited, but all these changes are not statistically significant. The CAT and LP activities both encountered an opposite trend of changes from SOD, and they increased in the first study and in the second one. The gross indices (CF, HSI) did not show any specific trend, HSI increased (not significantly), whereas CF decreased (not significantly) for the highest tested concentration of the mix of PAHs. The only statistically significant (po0.05) change in CF was its decrease for the concentration 0.148 mM in the first run. The most consistent results were obtained for FACs: Phe-, Pyr- and Fluometabolites experienced strong increase with the increase in the concentration of the mix of PAHs in the water.

4. Discussion Based on our previous studies on the biochemical response of gilthead seabream to selected individual PAHs: Phe, Pyr and Fluo (Correia et al., 2007; Kopecka-Pilarczyk and Correia, submitted), we conducted experiments with a mixture of these three PAHs, and in this paper we analyse the biomarker responses in seabream

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to the exposure to selected concentrations of this mixture of PAHs. Polycyclic aromatic hydrocarbons are metabolised primarily in Phase I by the cytochrome P450 (CYP) system, and also by several Phase II enzymes (Pretti et al., 2001), measured as EROD and GST, respectively. Then PAH metabolites are excreted into the bile and accumulate there (Aas et al., 1998), and are measured by fluorescence methods as FACs. EROD induction can be documented in fish exposed to spilled petroleum despite low observed tissue levels of PAHs (Whyte et al., 2000). However, in our study EROD decreased for the highest PAH mix concentration. This effect was similar to the hepatic EROD response to individual PAHs found in our previous studies: Phe (Correia et al., 2007), Pyr and Fluo (Kopecka-Pilarczyk and Correia, submitted). These results agree with the situation reported by Bucheli and Fent (1995) who noted that the activity of cytochrome CYP1A could be inhibited at higher concentrations of some inducers, such as b-naphtoflavone (bNF) and certain PCBs. On the other hand, the measured level of FACs consistently increased in our experiments with the increase in the concentration of the PAH mix. This decrease in EROD activity in the first experiment and increase in the level of PAH metabolites in bile, mainly for the highest tested concentration of the pollutants, indicate that the organisms increased PAH biotransformation. However, as reported by Gagnon and Holdway (2000), low molecular weight PAHs such as the two- and three-ring PAHs (Phe and Fluo are three-ring PAHs) are not known to be potent EROD activity inducers, as opposed to PAHs with four (like Pyr) or more condensed benzene rings, which are also known to have mutagenic or carcinogenic properties. When fish are exposed to PAHs, positive correlation between the induction of the mixedfunction oxygenase (MFO) system and the occurrence of PAH metabolites in the bile is expected. Indeed, as reviewed by Barra et al. (2001), such a relationship has been reported in a number of studies. For the two lower concentrations of the mix of PAHs tested in our study, the obtained EROD results are very variable, and even differ for the same retested concentration. This is similar to what was observed in our previous study with individual PAHs, so we are inclined to believe that the most probable reason for these variations and for the lack of a clear trend in the results is too short exposure time used in the experiments. The GST response to the tested mix of PAHs is similar to the response measured for Phe (Correia et al., 2007) and Fluo (Kopecka-Pilarczyk and Correia, submitted), which shows that fish detoxify the chemicals (Wang et al., 2006). In general, studies on Phase II inductive responses are less extensive than those concerned with Phase I activity, as reviewed by Gravato and Santos (2002) and van der Oost et al. (2003). Nevertheless, it has been determined that GST activity is influenced by the levels of organic substrates, and both induction and inhibition of these enzymatic activities have been reported in field polluted organisms (Regoli and Principato, 1995). The trend in BTI found in our experiments was similar to the trend observed for individual PAHs, mainly for the highest concentrations. This can be an indication of the disappearance of pollutants in the fish as they were metabolised (KopeckaPilarczyk and Correia, submitted). Regarding antioxidants for the mix of PAHs, they revealed a similar trend to the one observed for individual PAHs: CAT and LP increased in response to the highest concentration (0.444 mM), in opposite to SOD, which decreased for that concentration. These variations, however, are not statistically significant. On the other hand, in our experiments both GST and CAT increased, and we found significantly negative correlation between LP and GST (0.63, 2nd study, concentration 0.222 mM). It is commonly known that observed induction of antioxidants can provide sensitive early warning signals of incipient oxidative stress, and

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Fig. 1. Relative average activity of hepatic biomarkers (EROD, GST, BTI, CAT, SOD, LP) and gross indices (HSI, CF) of juvenile seabream exposed to the mix of PAHs (0.148, 0.222 and 0.444 mM). Error bars show the means with the 95% confidence belts, with K indicating the first, and J the second study.

oxidant-mediated toxicities may happen if these enzymes are inhibited, or enzymatic activation may occur if the system is impaired (Jee and Kang, 2005). Antioxidants may also protect tissues against oxidative damage caused by lipid peroxidation. Therefore, a possible explanation of our results is the tissue damage for the highest PAH mix concentration, where SOD decreased and LP increased. Failure to detoxify and remove reactive oxygen species (ROS) can result in oxidative damage to key molecules, including lipids and DNA, which in turn has been linked to disease processes, including cancer (as reviewed by Livingstone and Nasci, 2000). The increase in SOD suggests that more oxyradicals are produced, and it may also explain the rise in the CAT activity (Otto et al., 1994). A number of studies report that the measurement of fish bile PAH metabolites by fluorescence methods is an even more reliable biological indicator of PAH exposure than other, more general biological indicators such as the measurement of liver enzymes’ activities (Barra et al., 2001). These results were also confirmed in

our experiments, as we observed a consistent increase in FACs in response to the increasing concentration of the PAH mix. It has been observed that exposure of fish to EROD-inducing compounds, e.g. PAHs, and their mixtures results in abnormalities in growth and health of organs or tissues (as reviewed by Whyte et al., 2000). In our experiment, the HSI increased insignificantly for the medium and highest concentrations of the PAH mix, in contrast to CF, which decreased for the highest concentration. This observation can indicate enhanced detoxification activities in response to the presence of PAHs (Pereira et al., 1993) together with worse fish health and physiological condition (Kopecka and Pempkowiak, 2008). However, Lange et al. (1992) observed a decrease in EROD activity accompanying higher HSI in dab from the North Sea, which was interpreted as weakened induction of the P450 system in the liver of dab. Also Whyte et al. (2000) reviewed that fish with enlarged livers may have greater capacity to metabolise accumulated contaminants, so that acute induction of the P450 forms is short-lived. Moreover, it is speculated that

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Acknowledgments This research was supported by the Portuguese Fundac- a˜o para a Cieˆncia e a Tecnologia (FCT), post doc Grant no. SFRH/ BPD/32125/2006 for Justyna Kopecka-Pilarczyk, and the project BIOTEST no. POCI/MAR/56964/2004, co-financed by FEDER ˆncia e Inovac- ˜ through Programa Operational Cie ao 2010. The authors express their gratitude to Renata Gonc- alves for conducting the technical part of the experiments, assistance in fish dissection, and FAC measurements in bile, to Prof. Lucia Guilhermino for making a fluorimeter available for the EROD measurements, and to Dr. Pawe" Pilarczyk for linguistic revision of the manuscript and valuable remarks. For the purpose of this paper, A.D. Correia designed the experiments, including the determination of the concentrations of the PAH mix. J. Kopecka-Pilarczyk collected all the samples (with Renata’s assistance), measured all biomarkers in the samples, conducted the statistical analysis of all the data, prepared tables and figures, and wrote the entire paper.

References

Fig. 2. Relative average level of biliary FAC metabolites of juvenile seabream exposed to the mix of PAHs (0.148, 0.222 and 0.444 mM). Error bars show the means with the 95% confidence belts, with K indicating the first, and J the second study.

fish may respond to increased demand for xenobiotic metabolism by enlarging the liver rather than by increasing the specific activity of detoxification enzymes. On the other hand, elevated level of HSI is probably an indicator of chronic rather than recent pollution exposure, which may explain the small increase in HSI observed in our short-term experiments.

5. Conclusions Our experiments with a mixture of selected PAHs arrive at similar conclusions to those obtained in previous studies (Correia et al., 2007; Kopecka-Pilarczyk and Correia, submitted). Since among the biomarkers analysed in our study only FACs showed positive correlations with the increase in the concentration of pollutants, they seem to be the most adequate indicator of exposure to this type of pollutants. Out of the other tested biomarkers, only the response of EROD to the highest concentration of the mix of PAHs was statistically significant. The other biomarkers, mainly antioxidants at the highest concentration, indicated tissue damage, which might suggest too high concentration of the tested PAHs. Since the observed changes in biomarkers are very variable for the remaining concentrations, our results suggest that further experiments with different concentrations of the PAH mix are necessary in order to clarify how the mix of PAHs affects the enzymatic systems in juvenile seabream, and to provide a more comprehensive overview of the response pattern.

Aas, E., Beyer, J., Goksøyr, A., 1998. PAH in fish bile detected by fixed wavelength fluorescence. Mar. Environ. Res. 46 (1–5), 225–228. Allen, J.I., Moore, M.N., 2004. Environmental prognostics: is the current use of biomarkers appropriate for environmental risk evaluation? Mar. Environ. Res. 58, 227–232. Anderson, J.W., Lee, R.F., 2006. Use of biomarkers in oil spill risk assessment in the marine environment—scholarly review. Hum. Ecol. Risk Assessment 12, 1192–1222. Barra, R., Sanchez-Hernandez, J.C., Orrego, R., Parra, O., Gavilan, J.F., 2001. Bioavailability of PAHs in the Biobio river (Chile): MFO activity and biliary fluorescence in juvenile Oncorhynchus mykiss. Chemosphere 45, 439–444. Bucheli, T.D., Fent, K., 1995. Induction of cytochrome P450 as a biomarker for environmental contamination in aquatic ecosystems. Crit. Rev. Environ. Sci. Technol. 25 (3), 201–268. Chetiyanukornkul, T., Toriba, A., Kameda, T., Tang, N., Hayakawa, K., 2006. Simultaneous determination of urinary hydroxylated metabolites of naphthalene, fluorene, phenanthrene, fluoranthene and pyrene as multiple biomarkers of exposure to polycyclic aromatic hydrocarbons. Anal. Bioanal. Chem. 386, 712–718. Claiborne, A., 1985. Catalase activity. In: Greenwald, R.A. (Ed.), Handbook of Methods for Oxygen Radical Research. CRC Press, Boca Raton, FL, pp. 283–284. Correia, A.D., Gonc- alves, R., Scholze, M., Ferreira, M., Reis-Henriques, M.A., 2007. Biochemical and behavioral responses in gilthead seabream (Sparus aurata) to phenanthrene. J. Exp. Mar. Biol. Ecol. 347, 109–122. Faust, M., Altenburger, R., Backhaus, T., Blanck, H., Boedeker, W., Gramatica, P., Hamer, V., Scholze, M., Vighi, M., Grimme, L.H., 2001. Predicting the joint algal toxicity of multi-component s-triazine mixtures at low-effect concentrations of individual toxicants. Aquat. Toxicol. 56, 13–32. Ferreira, M., Moradas-Ferreira, P., Reis-Henriques, M.A., 2005. Oxidative stress biomarkers in two resident species mullet (Mugil cephalus) and flounder (Platichthys flesus) from a polluted site in river Douro estuary, Portugal. Aquat. Toxicol. 71, 39–48. Frasco, M.F., Guilhermino, L., 2002. Effects of dimethoate and beta-naphthoflavone on selected biomarkers of Poecilia reticulata. Fish Physiol. Biochem. 26, 149–156. Gagnon, M.M., Holdway, D.A., 2000. EROD induction and biliary metabolite excretion following exposure to the water accommodated fraction of crude oil and chemically dispersed crude oil. Arch. Environ. Contam. Toxicol. 38, 70–77. Gonc- alves, R., Scholze, M., Ferreira, A.M., Martins, M., Correia, A.D., 2008. The joint effect of polycyclic aromatic hydrocarbons on fish behaviour. Environ. Res. 108, 205–213. Gravato, C., Santos, M.A., 2002. Liver phase I and phase II enzymatic induction and genotoxic responses of b-naphthoflavone water-exposed sea bass. Ecotoxicol. Environ. Saf. 52, 62–68. Habig, W.H., Pabst, M.J., Jakoby, W.B., 1974. Glutathione-S-transferases. The first enzymatic step in mercapturic acid formation. J. Biol. Chem. 249 (22), 7130–7139. Haux, C., Fo¨rlin, L., 1988. Biochemical methods for detecting effects of contaminants on fish. Ambio 17 (6), 376–380. Jee, J.-H., Kang, J.-C., 2005. Biochemical changes of enzymatic defense system after phenanthrene exposure in olive flounder, Paralichthys olivaceus. Physiol. Res. 54, 585–591. Kopecka, J., Pempkowiak, J., 2008. Temporal and spatial variations of selected biomarker activities in flounder (Platichthys flesus) collected in the Baltic Proper. Ecotoxicol. Environ. Saf. 70 (3), 379–391.

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Kopecka-Pilarczyk, J., Correia, A.D. Biochemical response in gilthead seabream (Sparus aurata) to in vivo exposure to pyrene and fluorene, Submitted for publication. Krahn, M.M., Ylitalo, G.M., Buzitis, J., Bolton, J.L., Wigren, C.A., Chan, S.L., Varanasi, U., 1993. Analyses for petroleum-related contaminants in marine fish and sediments following the Gulf oil spill. Mar. Pollut. Bull. 27, 285–292. Lange, U., Jedamski-Grymlas, J., Siebers, D., Karbe, L., 1992. Ethoxyresorufin-Odeethylase and cytochrome P450 in the Liver of Dab (Limanda limanda (L.)) from the Central and Southern North Sea. Mar. Pollut. Bull. 24 (9), 446–451. Legeay, A., Achard-Joris, M., Baudrimont, M., Massabuau, J.-Ch., Bourdineaud, J.-P., 2005. Impact of cadmium contamination and oxygenation levels on biochemical responses in the Asiatic clam Corbicula fluminea. Aquat. Toxicol. 74, 242–253. Livingstone, D.R., Nasci, C., 2000. Biotransformation and antioxidant enzymes as potential biomarkers of contaminant exposure in goby (Zosterissor ophiocephalus) and mussel (Mytilus galloprovincialis) from the Venice Lagoon. In: Lasserre, P., Marzollo, A. (Eds.), The Venice Lagoon Ecosystem, Man and the Biosphere Series. The Parthenon Publishing Group 25, pp. 357–373. Lowry, O.H., Rosebrough, N.J., Farr, A.L., Randall, R.J., 1951. Protein measurement with the folin phenol reagent. J. Biol. Chem. 193, 265–275. McCarthy, J.F., Shugart, L.R., 1990. Biological markers of environmental contamination. In: McCarthy, J.F., Shugart, L.R. (Eds.), Biomarkers of Environmental Contamination. Lewis Publisher, Boca Raton, FL, pp. 3–16. McCord, J.M., Fridovich, I., 1969. Superoxide dismutase: an enzymatic function for erythrocuprein (hemocuprein). J. Biol. Chem. 244, 6049–6055. Meador, J.P., Stein, J.E., Reichert, W.L., Varanasi, U., 1995. Bioaccumulation of polycyclic aromatic hydrocarbons by marine organisms. Rev. Environ. Contam. Toxicol. 143, 79–165. Niki, E., 2000. Lipid peroxides. In: Taniguchi, N., Gutteridge, J.M.C. (Eds.), Experimental Protocols for Reactive Oxygen and Nitrogen Species. Oxford University Press, Oxford, pp. 156–160. Otto, D.M.E., Linsdtro¨m-Seppa¨, P., Sen, C.K., 1994. Cytochrome P450-dependent enzymes and oxidant-mediated responses in rainbow trout exposed to contaminated sediments. Ecotoxicol. Environ. Saf. 27, 265–280. Peakall, D., 1992. Animal Biomarkers as Pollution Indicators. Ecotoxicol. Ser. 1. Chapman & Hall, London, 291pp. Pereira, J.J., Mercaldo-Allen, R., Kuropat, C., Luedke, D., Sennefelder, G., 1993. Effect of cadmium accumulation on serum vitellogenin levels and hepatosomatic and gonadosomatic indices of winter flounder (Pleuronectes americanus). Arch. Environ. Contam. Toxicol. 24, 259–266.

Pretti, C., Salvetti, A., Longo, V., Giorgi, M., Gervasi, P.G., 2001. Effects of b-naphthoflavone on the cytochrome P450 system, and phase II enzymes in gilthead seabream (Sparus aurata). Comp. Biochem. Physiol. C 130, 133–144. Regoli, F., Principato, G., 1995. Glutathione, glutathione-dependent and antioxidant enzymes in mussel, Mytilus galloprovincialis, exposed to metals in different field and laboratory conditions: implications for a proper use of biochemical biomarkers. Aquat. Toxicol. 31, 143–164. Santodonato, J., Howard, P., Basu, D., 1981. Health and ecological assessment of polynuclear aromatic hydrocarbons. J. Environ. Pathol. Toxicol. 5, 1–364. Sole´, M., Porte, C., Barcelo´, D., 2000. Vitellogenin induction and other biochemical responses in carp Cyprinus carpio after experimental injection with 17aethynylestradiol. Arch. Environ. Contam. Toxicol. 38, 494–500. Stegeman, J.J., Brouwer, M., Di Giulio, R.T., Fo¨rlin, L., Fowler, B.A., Sanders, B.M., Van Veld, P.A., 1992. Enzyme and protein synthesis as indicator of contaminant exposure and effect. In: Huggett, et al. (Eds.), Biomarkers—Biochemical, Physiological and Histological Markers of Anthropogenic Stress. SETAC Special Publication Series, Lewis Publisher, pp. 235–334. US EPA (US Environmental Protection Agency), 1987. Health and environmental effects profile for pyrene. Prepared by the Environmental Criteria and Assessment Office, Office of Health and Environmental Assessment, US Environmental Protection Agency, Cincinnati, OH, for the Office of Solid Waste and Emergency Response. ECAO-CIN-P277. van der Oost, R., Lopes, S.C.C., Komen, H., Satumalay, K., van den Bos, R., Heida, H., Vermeulen, N.P.E., 1998. Assessment of environmental quality and inland water pollution using biomarker responses in caged carp (Cyprinus carpio): use of a bioactivation:detoxication ratio as a biotransformation index (BTI). Mar. Environ. Res. 46 (1–5), 315–319. van der Oost, R., Beyer, J., Vermeulen, N.P.E., 2003. Fish bioaccumulation and biomarkers in environmental risk assessment: a review. Environ. Toxicol. Pharm. 13, 57–149. Wang, Ch., Zhao, Y., Zheng, R., Ding, X., Wei, W., Zuo, Z., Chen, Y., 2006. Effects of tributyltin, benzo[a]pyrene, and their mixture on antioxidant defense systems in Sebastiscus marmoratus. Ecotoxicol. Environ. Saf. 65, 381–387. Whyte, J.J., Jung, R.E., Schmitt, C.J., Tillitt, D.E., 2000. Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit. Rev. Environ. Sci. Technol. 30 (4), 347–570. Winston, G.W., Di Giulio, R.T., 1991. Prooxidant and antioxidant mechanisms in aquatic organisms. Aquat. Toxicol. 19, 137–161.