14
Biochemical and behavioral responses in gilthead seabream (Sparus aurata) to phenanthrene Ana D. Correia a,b, , Renata Gonçalves a , Martin Scholze c , Marta Ferreira a,d , Maria Armanda-Reis Henriques a,d a CIIMAR-Centro de Investigação Marinha e Ambiental, Laboratório de Toxicologia Ambiental, Porto, Portugal b Instituto de Biopatologia Química, Faculdade de Medicina de Lisboa, Unidade de Biopatologia Vascular, Instituto de Medicina Molecular, Lisboa, Portugal c The School of Pharmacy, University of London, London, United Kingdom d ICBAS-Instituto de Ciências Biomédicas de Abel Salazar, Porto, Portugal Received 11 December 2006; received in revised form 21 March 2007; accepted 27 March 2007 Abstract Most toxicological studies with PAHs investigate their impact on aquatic organisms only at very specific levels of organization, either at molecular and cellular levels via biomarkers, or at higher integral endpoints such as reproduction and behavior. The link between both has received less attention in science. The aim of this multi-response study was to investigate the relationship between specific molecular processes (induction of biotransformation enzymes and oxidative stress) and the behavioral performance of fish. We performed two concentration-effect studies with juvenile gilthead seabream (Sparus aurata), at which fish were exposed for 4 days to phenanthrene (PHE) (0.11 to 0.56 μM). Groups of five fish per aquarium were recorded for changes in the patterns of their movement and social interactions. Biomarkers analyzed were ethoxyresorufin-O-deethlylase (EROD), total glutathione-S-transferase (GST), phenanthrene-type metabolites in bile, lipid peroxidation (LP), superoxide dismutase (SOD) and catalase (CAT). The physiological status of the fish was determined by the liver somatic index. In general, PHE changed the overall behavioral performance of fish, all behavior activities were affected in a dose-response way. The incidence of lethargic fish was strongly increased (up to 39%), as the fish activities were reduced. The changes in the individual swimming activity had influenced negatively the social behavior of fish groups, i.e. the more fish in the group were lethargic, the less the social interactions were marked. The biomarkers responded to PHE differently, with an increase of EROD activity at low exposures (72.25 pmol min 1 prot 1 ), but an inhibition at high concentrations (42.60 pmol min 1 prot 1 ). For GST, we observed the reverse pattern. Together with the strong increase of PHE-type metabolites in bile, we conclude that both biotransformation enzymes are involved in the metabolism of PHE in liver. We found indications for oxidative stress already at low PHE concentrations, as LP levels were increased in the liver. However, higher exposures provoked less pronounced levels, but elevated activities of the antioxidants CAT and SOD (up to 37% and 17%, respectively). We conclude that especially the enzymatic activations at high-PHE exposures might have required additional energetic costs for the chemical detoxication that lead to the marked changes in the fish behaviors, i.e. demonstrating a trade-offbetween detoxication processes via the biliaryhepatic system and the fish activity. Thus, the strong increases in lethargy might be the Journal of Experimental Marine Biology and Ecology 347 (2007) 109 122 www.elsevier.com/locate/jembe Corresponding author. CIIMAR-Centre of Marine and Environmental Research, Environmental Toxicology Laboratory, Rua dos Bragas, 289, 4050-123 Porto, Portugal. Tel.: +351 223401833; fax: +351 223390608. E-mail address: [email protected] (A.D. Correia). 0022-0981/$ - see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.jembe.2007.03.015

Biochemical and behavioral responses in gilthead seabream (Sparus aurata) to phenanthrene

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y and Ecology 347 (2007) 109–122www.elsevier.com/locate/jembe

Journal of Experimental Marine Biolog

Biochemical and behavioral responses in gilthead seabream(Sparus aurata) to phenanthrene

Ana D. Correia a,b,⁎, Renata Gonçalves a, Martin Scholze c,Marta Ferreira a,d, Maria Armanda-Reis Henriques a,d

a CIIMAR-Centro de Investigação Marinha e Ambiental, Laboratório de Toxicologia Ambiental, Porto, Portugalb Instituto de Biopatologia Química, Faculdade de Medicina de Lisboa, Unidade de Biopatologia Vascular,

Instituto de Medicina Molecular, Lisboa, Portugalc The School of Pharmacy, University of London, London, United Kingdomd ICBAS-Instituto de Ciências Biomédicas de Abel Salazar, Porto, Portugal

Received 11 December 2006; received in revised form 21 March 2007; accepted 27 March 2007

Abstract

Most toxicological studies with PAHs investigate their impact on aquatic organisms only at very specific levels of organization,either at molecular and cellular levels via biomarkers, or at higher integral endpoints such as reproduction and behavior. The linkbetween both has received less attention in science. The aim of this multi-response study was to investigate the relationship betweenspecific molecular processes (induction of biotransformation enzymes and oxidative stress) and the behavioral performance of fish.We performed two concentration-effect studies with juvenile gilthead seabream (Sparus aurata), at which fish were exposed for4 days to phenanthrene (PHE) (0.11 to 0.56 μM). Groups of five fish per aquarium were recorded for changes in the patterns of theirmovement and social interactions. Biomarkers analyzed were ethoxyresorufin-O-deethlylase (EROD), total glutathione-S-transferase(GST), phenanthrene-type metabolites in bile, lipid peroxidation (LP), superoxide dismutase (SOD) and catalase (CAT). Thephysiological status of the fish was determined by the liver somatic index. In general, PHE changed the overall behavioralperformance of fish, all behavior activities were affected in a dose-response way. The incidence of lethargic fish was stronglyincreased (up to 39%), as the fish activities were reduced. The changes in the individual swimming activity had influenced negativelythe social behavior of fish groups, i.e. the more fish in the group were lethargic, the less the social interactions were marked. Thebiomarkers responded to PHE differently, with an increase of EROD activity at low exposures (72.25 pmol min−1 prot−1), but aninhibition at high concentrations (42.60 pmol min−1 prot−1). For GST, we observed the reverse pattern. Together with the strongincrease of PHE-type metabolites in bile, we conclude that both biotransformation enzymes are involved in the metabolism of PHE inliver. We found indications for oxidative stress already at low PHE concentrations, as LP levels were increased in the liver. However,higher exposures provoked less pronounced levels, but elevated activities of the antioxidants CAT and SOD (up to 37% and 17%,respectively). We conclude that especially the enzymatic activations at high-PHE exposures might have required additional energeticcosts for the chemical detoxication that lead to the marked changes in the fish behaviors, i.e. demonstrating a “trade-off” betweendetoxication processes via the biliary–hepatic system and the fish activity. Thus, the strong increases in lethargy might be the

⁎ Corresponding author. CIIMAR-Centre of Marine and Environmental Research, Environmental Toxicology Laboratory, Rua dos Bragas, 289,4050-123 Porto, Portugal. Tel.: +351 223401833; fax: +351 223390608.

E-mail address: [email protected] (A.D. Correia).

0022-0981/$ - see front matter © 2007 Elsevier B.V. All rights reserved.doi:10.1016/j.jembe.2007.03.015

110 A.D. Correia et al. / Journal of Experimental Marine Biology and Ecology 347 (2007) 109–122

consequence of higher energetic demands for the PHE detoxication. This illustrates how an integrated use of biomarkers cancontribute to our understanding of the impact of PAHs at increasing levels of biological complexity.© 2007 Elsevier B.V. All rights reserved.

Keywords: Behavior endpoints; Biomarkers; Metabolism; Phenanthrene; Seabream; Water exposures

1. Introduction

Organic xenobiotics in aquatic ecosystems originatemainly from the production of synthetic chemicals andthe use of fossil-energy. Of particular interest are poly-cyclic aromatic compounds (PAHs), a group of over 100different chemicals that are formed during the incom-plete burning of coal, oil and gas, garbage, or otherorganic substances like tobacco or charbroiled meat. In2004, their quantities were estimated to exceed 2 millionpounds in the US (US EPA, 2006). Organic xenobioticsare a potential threat to humans and the environment,especially with respect to PAHs which are suspected tobe carcinogens (Albers, 2003). Because of their abilityto absorb easily to organic materials (Law and Biscaya,1994) they are commonly found as pollutants in soils,estuarine waters and sediments, and other terrestrial andaquatic sites.

Most toxicological studies with PAHs have beeninvestigated at molecular and cellular levels. For in-stance, enzymes that are part of the [Ah]-gene battery(e.g. CYP1A, enzymes of phase II conjugates, antiox-idant enzymes) are often used as molecular biomarkersin order to investigate the influence of PAHs on thebiochemical pathways and enzyme functioning in fish(reviewed byWhyte et al., 2000), and many studies havedemonstrated that fish possess a well-developed MFOsystem that might efficiently detoxify a large number ofxenobiotics, including PAHs. During the processes ofdetoxication, reactive metabolites can be produced andelicit toxicity through the generation of reactive oxy-gen species (ROS) and/or for binding covalently to cel-lular macromolecules such as DNA, RNA and protein(reviewed by Van der Oost et al., 2003).

Biomarkers at molecular level are meaningful be-cause they are able to respond quickly and often highlyspecific to chemical stressors (Van der Oost et al., 2003).However, their value is limited when we want to assessthe impact of exposures for the whole organism, mainlyas the link between biochemical responses and higherintegral endpoints such as physiology, reproduction andbehavior are too often unclear (Jensen et al., 1997;Livingstone, 2001). Studies looking on how the differentlevels of biological organization are related to each otherthus improve the mechanistic understanding of toxicity

and their ecological consequences (Weis et al., 2001).Although it has been shown that PAHs can interfere onhigher levels of organization in fish (e.g. Farr et al., 1995;Monteiro et al., 2000; Jee et al., 2004), the majority of thestudies have investigated the impact of these compoundsonly at very specific levels of organization.

Individual behavior is an integral response parameterthat is linked to activities at biochemical levels, e.g.changes in the swimming activity of fish are commonlythe result of damages in the nervous and hormonalcontrol system, induced by metals and polychlorinatedbiphenyls—PCBs (Jensen et al., 1997,Weis et al., 2001).Although PAHs can affect the behavior performance offish (Westlake et al., 1983; Farr et al., 1995; Hinkle-Connet al., 1998), it remains unknown how this can be linkedto disruptions at biochemical level. Typically, reproduc-tion endpoints are used to assess chemical effects onpopulation and community levels, but individual changesin behavior can provide similar information (Weis et al.,2001), as toxicants can disturb behavioral patterns thatare essential for the fitness and survival of the entirepopulation (Scott and Sloman, 2004). Thus, behavioralendpoints and their mechanistic understanding are animportant step to analyze the connections between subtlebiochemical changes in the organism and their ecologicalconsequences.

Phenanthrene (PHE) is a priority PAH, and, althoughnot mutagenic or carcinogenic, it has been shown to betoxic to marine diatoms, gastropods, mussels, crusta-ceans, and fish (Albers, 2003; US EPA, 2006). SincePHE is the smallest tricyclic aromatic hydrocarbon tohave a “bay-region” and a “K-region” (Ouyang, 2006),i.e. highly reactive regions of PAH molecules where themain carcinogenic species can be formed, it is common-ly used as a model substrate for studies on metabolismof carcinogenic PAHs.

We used PHE as a model compound in order toinvestigate the relationship between specific molecularprocesses (induction of biotransformation enzymesand oxidative stress) and the behavioral performance offish at the individual level. Juvenile seabream (Sparusaurata) was used asmodel species because of its ability toyield reproducible behavior data under controlled condi-tions (Begout and Lagardere, 1995). Seabream is widelycultured in Europe. We performed two concentration-

111A.D. Correia et al. / Journal of Experimental Marine Biology and Ecology 347 (2007) 109–122

effect studies with PHE in a semi-static system, at whichfish were exposed daily for 4 days. At the end, liveractivities of ethoxyresorufin-O-deethlylase (EROD), totalglutathione-S-transferase (GST) and PHE-type metabo-lites in bile were analyzed in order to gain insight into thedetoxication and excretion mechanisms for PHE. Fur-thermore, we measured the responses of catalase (CAT)and superoxide dismutase (SOD) as indicators of oxida-tive stress and levels of lipid peroxidation (LP) as indi-cation of damage. The physiological status of the fish wasdetermined by the liver somatic index (LSI). We exam-ined the behavior by recording visually the activity andsocial interactions of groups of fish per aquarium.

2. Material and methods

2.1. Test organisms and chemicals

Gilthead seabream juveniles, S. aurata, L., weresupplied from a commercial fish farm (TIMAR Lda.,Setúbal, Portugal), where they had been raised till theweight of 1.0 g. All fish were from the same batch, andbefore dosing, they were kept under laboratory condi-tions in 60-l aquaria (density 2–3 g m−3) supplied withfiltered seawater (35 ±2 ppm). The fish were fed dailywith a maintenance ration of 2–3% body weight, andtheir average body weight during the exposures were2.0 ±0.2 g (first study, n =75) and 2.2 ±0.2 g (secondstudy, n =75). PHE (≥ 97% purity) was purchasedfrom Aldrich (Milwankee, WI). All other chemicalswere of analytical grade and obtained from Sigma(St. Louis, USA), and E. Merck (Darmstadt, Germany).

2.2. Experimental design

Waterborne exposures were conducted in 17-l glassaquaria at 16 ±1 °C in filtered seawater (35 ±2 ppm)under a photoperiod of 12 h light: 12 h dark. The aquariawere kept at semi-obscurity during the light periods inorder to avoid PHE phototoxicity. Dissolved oxygensaturation (N 80%) and total ammonia concentrations(b 0.5 mg l−1) were monitored weekly. Aquaria wereconstructed of glass, and the contact of other materials(e.g. silicon rubber tubing) with the test solutions wasminimized. PHE was initially dissolved in acetone, andthe stock solution was kept at −20 °C until prepared forthe final exposure solutions in seawater. Exposures weredaily renewed along with seawater (50% of total volume),and the solvent concentrations never exceeded 0.0014%in the aquaria. Water disposal from the aquaria wasfiltered through activated carbon before being deliveredinto the municipal sewage system.

Before exposure, animals were acclimatized at thesame conditions described for waterborne exposures for24 h in 20-l aquaria and then five randomly chosen fishwere placed in each test aquarium for 24 h with aeration(pre-exposure phase). Afterwards, fish were exposed toPHE for 4 days (post-exposure) (ASTM, 2003). Foodwas not provided during the acclimation and in thecourse of exposures. Aeration was provided with plastictips placed 2 cm above the aquaria bottom.

We conducted two consecutive studies within 1month,at nominal concentrations of 0.11 and 0.56 μM PHE inthe first study (0.02 and 0.1 mg l−1, respectively), and0.11 and 0.28 μM in the second (0.02 and 0.05 mg l−1,respectively). Five fish per aquarium were used, andin order to account for the inter-aquarium variability,always five aquaria per treatment and control (acetone).We recorded daily the individual behavior of the fish,starting 1 day before the exposure begins (day 0). After4 days of exposure, the animals were sacrificed for thesubcellular analyses.

2.3. Sample preparation

For the biochemical analysis, we always pooled livertissues (50–100 mg wet weight) from two fish. Liverswere homogenized in ten volumes of phosphate buffer(100 mM, pH 7.5) containing 1 mM EDTA, and after-wards centrifuged at 10,000 ×g for 20 min at 4 °C. Wedistributed the resulting postmitochondrial supernatants(PMS) into aliquots and stored them at −80 °C prior toanalysis. We used a sample volume of 250 μl for thelipid peroxidation assay and of 40 μl for the enzymaticmeasurements (EROD, GST, CAT, and SOD). Totalprotein concentration was determined in PMS super-natants according to the Lowry method (Lowry et al.,1951) and adapted to microplates using bovine serumalbumin as standard. We collected bile samples byincising the gall bladder and stored them at −80 °C untilthe analysis. We excised and weighed individual liversand determined the liver somatic index (LSI) as thepercentage ratio of liver weight to body weight.

2.4. Biochemical assays

EROD activity was measured by the fluorimetricmethod described in Solé et al. (2000). PMS liversamples (25 μl) were incubated at 30 °C for 10 min in afinal volume of 0.5 ml, containing phosphate buffer(87 mM, pH 7.5), 0.22 mM NADPH, and 3.70 μM 7-ethoxyresorufin. The reaction was stopped by adding1 ml of ice-cold acetone, samples were centrifuged at400 ×g and 7-hydroethoxyresorufin fluorescence was

112 A.D. Correia et al. / Journal of Experimental Marine Biology and Ecology 347 (2007) 109–122

determined at 530/585 nm excitation/emission wave-lengths. We expressed the EROD activity as pmol min−1

mg prot−1.Total glutathione-S-transferase (GST) was determined

using 1-chloro-2,4-dinitrobenzene (CDNB) according tothe method of Habig et al. (1974) and adapted tomicroplates (Frasco and Guilhermino, 2002). Reactionmixtures contained 4.95 ml phosphate buffer (0.1 M, atpH 6.5):0.9 ml GSH (10 mM):0.15 ml CDNB (60 mM).In the microplate, we added 0.2 ml of the reaction mix-ture to 0.1 ml of the sample, with a final concentrationof 1 mM GSH and 1 mM CDNB in the assay. GSTwas measured using CDNB as change in OD/min at340 nm (ε =9.6 mM−1 cm−1) and expressed as nmolmin−1 mg prot−1.

Catalase (CAT) activity wasmeasured by the decreasein absorbance at 240 nm because of H2O2 consumption(ε =40 M−1 cm−1). For the reaction, we used 67.5 mMpotassium phosphate buffer (pH 7.5) and 12.5 mMH2O2, and initiated it with the addition of the sample. Weexpressed the CAT as μmol min−1 mg prot−1.

SOD activity was determined in the PMS liverfraction as inhibition of cytochrome c reduction at550 nm (McCord and Fridovich, 1969), adopted tomicroplate (Ferreira et al., 2005). The reaction containedphosphate buffer (50 mM, pH 7.8), 50 μM hypoxan-thine, 1.98 mU ml−1 xanthine oxidase and 10 μM cyto-chrome c. We measured the relative activity in unitsof SOD (U mg prot−1), with one unit SOD being theamount of sample causing a 50% inhibition of cyto-chrome c reduction under the standard conditions ofthe assay.

Tissue lipid peroxides (malondialdehyde—MDAequivalents) were measured in PMS by the thiobarbi-turic acid method (Niki, 2000). Subsamples of tissuehomogenate were incubated with 100% TCA, and aftercentrifugation the supernatant was incubated for 30 minat 100 °C with 1% TBA, 0.05 M NaOH and 0.025%BHT. The supernatant (organic layer) was taken and itsabsorbance measured at 532 nm (ε =1.54 × 105 M−1

cm−1, Halliwell and Gutteridge, 1999). We expressedthe lipid peroxidation (LP) as MDA equivalents per mgliver (wt.).

2.5. Phenanthrene-type metabolites analysis

We diluted the bile samples from controls in 48%ethanol to 1:1500, and samples from exposed bile to1:100,000. Fluorescent readings were made at 260/380 nm (excitation/emission) for PHE-type metabolites(Krahn et al., 1993) using PHE as a reference standard.We used a 5-nm slit width for excitation and emission.

2.6. Behavioral assays

Fish were randomly assigned to the treatments andthe same person recorded their behavior in all studies. Inorder to avoid a recording bias, exposures wereunknown to this person. Individual fish responseswere monitored each day for 2 h from 10 to 12 a.m.by recording visually at every 12 min their behavioralactivity and spatial distribution as an “all-or-none”response, obtaining 10 counts for each aquarium persession. After recording, aqueous PHE and solventcontrol solutions were administered to the aquaria. Thestudy was completed after 5 days, with one pre- and fourpost-exposure data sets of behavioral records for eachaquarium. Check sheets were used to record behavioralobservations. The basic design of the check sheet was agrid, with columns denoting successive sample intervalsand rows denoting the behavior endpoints defined(Martin and Bateson, 1993). Each individual behavioralactivity was categorized into three types: (i) swimmingpatterns, defined in terms of horizontal movements(swimming) and vertical movements (rising), (ii)lethargy, described as a non-locomotory activity by theabsence of detectable body movements, and (iii) socialpatterns (social interaction), i.e. avoiding, biting orchasing behavior. Such behavioral categories can beaffected by environmental contaminants, includingPAHs (Sorensen et al., 1997; Sloman et al., 2003).Additionally, the position of each fish in the aquaria wasrecorded (bottom, middle, surface and near the aerationfilter) (Yilmaz et al., 2004).

2.7. Statistical analyses

Biomarker data were examined for normal distribu-tion and homogeneity of variance (Shapiro–Wilk's andBartlett's tests), and if required, data were log trans-formed. Dunnett's test (α =5%) was then employed todetermine whether any of the treatment groups differedin relation to the solvent controls, with aquarium alwaysa nested factor in data analysis.

We proved that the behavioral count data follow anoverdispersed binomial distribution, which lead todifficulties for the data analysis: As the experimentaldesign is nested, and differences between control andexposure means are of main interest, it formulates anunsolved problem in statistics and ruled out a powerfuldata analysis. Instead, we used the aquaria means asstatistical units and assessed the differences betweencontrols and exposures by the nonparametric Kruskal–Wallis test. A further quantitative difficulty was thecorrect choice of the control reference for the post-

Table1

Meaneffectsforphenanthrene

exposuresto

juvenile

seabream

(alwaysnested

analysiswith

factor

aquarium

,data

alwayslogtransformed)

End

point

Dun

nett

Firststud

ySecon

dstud

y

Control

0.11

μM0.56

μMReplicates

Control

0.11

μM0.28

μMReplicates

EROD

(pmol

min−1prot−1)

Two-sided

71.1

104.0

42.6⁎

8–10

46.85

72.25⁎

52.64

6–9

[59.9;84

.2]

[79.7;13

5.7]

[29.9;60

.7]

[38.29

;57.33

][48.65

;107

.30]

[37.33

;74.24]

GST(nmol

min−1prot−1)

Two-sided

143.5

107.9

160.2

1014

0.7

124.8

133.9

10[113

.3;181

.7]

[87.6;13

2.7]

[138.1;185

.9]

[107.5;184

.2]

[108.3;143

.8]

[117

.7;152

.4]

CAT(µmol

min−1prot−1)

One-sided

45.2

51.5

57.9⁎

842

.051

.9⁎

49.2⁎

9[37.0;55

.3]

[44.9;59

.1]

[52.2;64

.3]

[39.2;45

.0]

[48.2;55

.9]

[47.2;51.2]

SOD

(Umgprot

−1)

One-sided

15.4

15.41

17.9

9–10

15.64

16.77

16.97

10[14.0;16

.9]

[14.79

;16.04

][15.1;21

.2]

[13.88

;17.63

][14.87

;18.92

][14.16

;20.33]

PL(nmol

MDA

g−1)

Two-sided

58.3

64.4

58.0

7–10

49.7

65.2

62.9

6–10

[49.0;69

.3]

[53.3;77

.9]

[37.8;88

.9]

[41.7;59

.3]

[42.8;99

.3]

[52.4;75.6]

LSI(%

)One-sided

1.34

1.46

1.61

⁎25

1.21

1.18

1.38

25[1.25;1.44]

[1.29;1.65]

[1.44;1.80

][1.11;1.33]

[1.10;1.28

][1.32;1.45]

Phenanthrene-type

metabolites(μgml−

1,pp

m)

One-sided

102.5

2728

⁎10

252⁎

13–15

151.0

3389

8928

⁎15

[87.5;12

0.0]

[2254;33

01]

[8324;12

625]

[110

.7;206

.1]

[2704;42

48]

[759

1;1050

1]

⁎Significant

atα=5%

.

113A.D. Correia et al. / Journal of Experimental Marine Biology and Ecology 347 (2007) 109–122

exposed fish, with typically two possible approaches. Oneapproach is to compute the difference of response valuesfor each fish group in a given aquarium (treatments andcontrols) with its own control values prior to exposure,and a mean difference is then estimated for a givenbehavioral response variable in each aquarium at a giventime (individual pre–post comparison). This mean differ-ence is then compared to the mean difference observed inthe control fish aquarium (ASTM, 2003). This approachassumes that a fish (or group) “remembers” its pre-exposure behavior. Alternatively, in the other approach,the post-exposure data for the treatments are comparedwith the post-control data (post–post comparison), whichrequires no pre-exposure information. Because of the longexposure duration and the absence of food for the juvenilefish, we found the latter approach also suitable and usedtherefore both approaches. All analyses were performedusing the SAS procedure PROC GENMOD and PROCGLM (SAS version 9, SAS Institute Inc, Cary, NC, USA).

3. Results

Two concentration-effect studies with juvenile gilt-head seabream (S. aurata) were performed, at which fishwere exposed for 4 days to phenanthrene (PHE) (0.11 to

Fig. 1. Relative average activity of ethoxyresorufin-O-deethlylase—EROD (pmol min−1 mg prot−1) (A) and total glutathione-S-transferase—GST (nmol min−1 mg prot−1) (B) in liver of juvenile seabreamexposed to phenanthrene (0.11, 0.28 and 0.56 μM). Data were re-scaled by subtracting the control mean from each measurement. Errorbars show the mean with 95% confidence belts, with ● from first and ○

from the second study.

Fig. 3. Concentration–response data of phenanthrene-type metabolites−1

114 A.D. Correia et al. / Journal of Experimental Marine Biology and Ecology 347 (2007) 109–122

0.56 μM). Both studies were conducted successfully,and we observed neither fish mortality nor anyevidences for fish infections or other diseases. Statisticalresults about the average enzymatic activities in juvenilefish are given in Table 1, together with the 95% confi-dence intervals, the number of replicates and allstatistical test decisions. In order to achieve better datacomparability between both studies, we re-scaled thesedata by subtracting the control mean from the dataobservations, and results are shown in Figs. 1 and 2.Outcomes from the analysis of PHE-type metabolites infish bile are pictured in Fig. 3. The behavioral activitiesof fish before and after 4 days of dosing are summarizedin Table 2, and the corresponding relative changes to thecontrols are presented in Table 3 and Fig. 4. In Fig. 5, the

Fig. 2. Relative average activity of catalase—CAT (μmol min−1 mgprot−1) (A), superoxide dismutase—SOD (U mg prot−1) (B) andaverage lipid peroxidation—malondialdehyde levels (nmol g−1 liverwt.) (C) in liver of juvenile seabream exposed to phenanthrene (0.11,0.28 and 0.56 μM). Data were re-scaled by subtracting the controlmean from each measurement. Error bars show the mean with 95%confidence belts, with ● from first and ○ from the second study.

(equivalents, μg ml ) in bile of juvenile seabream exposed tophenanthrene (0.11, 0.28 and 0.56 μM). The black dots represent themeans, connected by a smoothing solid line.

relationships between observed behavioral changes andmeasured enzyme activities are pictured exemplarily foreight selected cases, i.e. two behavioral parameters(lethargy and social interactions) are related to EROD,GST, CAT and SOD responses.

3.1. Subcellular responses

Pooled data for EROD activity in seabream showed anon-monotonic concentration–response relationship(Fig. 1A), as for the lowest tested concentration(0.11 μM PHE) the measurements were about 1.5-foldhigher than control values, while at the highest testedconcentration (0.56 μM PHE) the activity was signif-icantly suppressed ( p b 0.05). Although we detectedthe observed increase as statistically significant only fordata from the second study, both studies yielded similarmean values. Indeed a pooled data analysis for the re-scaled data (with study as co-factor in the model)confirmed the statistical significance (data not shown).

Concentration–response data for GST activity indi-cate the opposite pattern: The measurements were inaverage 25% lower for 0.11 μM PHE, but 14% higherfor 0.56 μM PHE (Fig. 1B). However, because of higherdata variation the statistical power was not sufficientlyhigh to detect these small differences as statisticallysignificant.

Both antioxidant enzymes, CAT and SOD, increasedwith higher PHE concentrations (Fig. 2A and B), withCAT levels in liver at the highest tested concentrationaround 30% higher than in controls (p b 0.05). However,SOD levels in exposed fish were much less enhanced,e.g. for 0.56 μM PHE in average only 17% higher levelswere measured.

Table 2Percentual fish activities of pre a and post b-exposed juvenile seabream to phenanthrene

First study Second study

Control 0.11 μM 0.56 μM Control 0.11 μM 0.28 μM

Pre Post Pre Post Pre Post Pre Post Pre Post Pre Post

Behavioral activitiesSwimming 24.8 29.6 30.4 23.7 20.4 12.3 46.8 41.2 49.6 34.8 46.0 28.4Rising 6.8 9.8 10.8 7.0 8.0 2.4 26.0 11.2 13.6 6.4 27.6 6.4Lethargy 34.9 29.6 34.0 38.3 41.6 68.2 12.4 19.2 17.2 23.6 12.8 42.4Social interactions 33.5 30.9 24.8 30.8 30.0 17.0 14.8 28.4 19.6 35.2 13.6 22.8

Position in the aquariaSurface 0.0 2.4 0.0 1.2 1.6 0.0 0.8 2.8 0.0 2.0 0.4 1.2Middle 88.2 88.1 77.6 82.3 76.4 87.2 88.4 82.8 93.6 95.2 94.8 92.8Bottom 11.8 7.1 21.6 11.4 21.6 9.2 10.8 14.4 6.4 2.8 4.8 6.0Aeration filter 0.0 2.4 0.8 5.1 0.4 3.6 0.0 0.0 0.0 0.0 0.0 0.0a Fish were recorded before dosing.b Fish were recorded after four-day exposures.

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Levels of liver lipid peroxide, measured in terms ofMDA, indicate a non-monotonic concentration–responsepattern similarly to that of EROD activity (Fig. 2C): Inboth studies, we measured highest levels of MDA at0.11 μM PHE, but observed less pronounced levels forhigher test concentrations. However, the increase was atmaximum only 8–10% above controls, and, with a coef-ficient variation of 20–30% in the controls, thus far belowthe minimal detection limit for statistics.

The analysis of fluorescent aromatic compounds(FACs) showed clearly the presence of PHE-typemetabolites in fish bile (Fig. 3), with average levels of0.1–0.15 μg ml−1 in the controls. Moreover, a clearconcentration-dependent accumulation was evident,and at 0.11 μM PHE already a 26-fold higher levelin bile fluid was measured (2.7 μg ml−1, p b 0.05). The

Table 3Behavioural changes in juvenile seabream after four-day exposures tophenanthrene

Control First study Second study

0.11 μM 0.56 μM 0.11 μM 0.28 μM

Behavioral activitiesSwimming 0 −5.9 −17.3 ⁎ −6.4 −12.8Rising 0 −2.8 −7.4 ⁎ −4.6 −4.8Lethargy 0 8.7 ⁎ 38.6 ⁎ 4.4 23.2 ⁎

Social interactions 0 −0.1 −13.9 6.8 −5.6

Position in the aquariaSurface 0 −1.2 −2.4 −0.8 −1.6Middle 0 −5.8 −0.9 12.4 ⁎ 10.0Bottom 0 4.3 2.1 −11.6 −8.4Aeration filter 0 2.7 1.2 0.0 0.0

All values are in percentages.⁎ Significant at α =5%.

concentration–response curve indicates that the accu-mulated levels in bile reached nearly a steady state at0.28 μM PHE, with higher exposure concentrationsproducing only a minor increase in fluorescence.

3.2. Liver somatic index

Compared to control fish, the PHE exposures pro-duced only slight changes in liver somatic index (LSI),and only the highest concentration at 0.56 μM PHEcaused a significant increase (Table 1).

3.3. Behavioral responses

We have summarized the recorded behavioral activ-ities for the tested juvenile fish in Table 2, for bothstudies, and always before and after exposures. Eachvalue represents the mean percentage activity from atotal of 25 fish, observed in five aquaria. Behavioralperformance parameter are categorized according to theswimming activities (swimming, rising, lethargy, andsocial interactions) and their position in the aquarium(surface, middle, bottom, and near aeration filter), i.e.mean values sum up to 100% for each.

In both studies, the majority of fish stayed during therecording period in themiddle of the aquarium (82–95%),and the fish avoided the surface and the proximity toaeration filter. However, the swimming activities differedbetween the studies: In the first study, around 70% of thecontrol fish showed behaviors like lethargy and socialinteractions at study begin, which were reduced to 60%after 4 days. However, in the second study these activitieswere reduced to 27.2% prior dosing and to 47.6% after4 days. The reason for these differences remains unclear.

Fig. 4. Average changes in lethargy (A), social interaction (B), swimming (C) and rising (D) in juvenile seabream exposed to phenanthrene (0.11,0.28 and 0.56 μM). The responses (%) are normalized to the solvent controls and based on nominal concentrations. Each ○ depicts the mean from anaquarium, each black dot (●) the overall mean from the first study and grey dot ( ) the overall mean from the second study. For a better visuality, datafrom second study are shifted slightly to the right.

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Furthermore, the differences in swimming activitiesbefore and after dosing were in the controls from thesecond study much more pronounced, probably becauseof the generally reduced non-locomotor activities in thesecond study. Although these differences between thestudies clearly deny a simple data pooling, both studiesnevertheless have produced very similar concentration–response pattern. This can be identified in the best waywhen not only the absolute concentration–responsevalues are compared, but also changes to average pre-or post-control activity are considered. Table 3 shows thechanges for post-exposure data when compared to theaverage control activity after 4 days, which are visualizedfor the swimming parameters in Fig. 4. It shows clearlythat PHE has caused a change in the overall performanceof the fish samples, as all behavior activities were affectedin a dose-response way. The relatively good agreementbetween the outcomes from both studies supports this,despite the considerable large inter-aquarium variability

that we observed for some of the selected endpoints(indicated by the scatter of small dots). When we base thechanges solely on a pre–post comparison (normalized tothe mean difference of the controls) then these concen-tration–response relationships were masked by a hugedata variation and changes are not anymore identifiable(data not shown). This indicates that the originalbehavioral pattern of fish in an individual aquarium wasnot maintained over the study duration, and consequentlythe pre-exposure information is not required for a mean-ingful concentration–response analysis.

The observed concentration–response pattern variedsubstantially between the locomotor and non-locomotoractivities. The clearest results were obtained for theincreased frequency of lethargic fish in the treatments,particularly at the two highest test lethargy concentrations(23.2% at 0.28 μM PHE and 38.6% at 0.56 μM PHE).For both, we tested the changes as statistically significant( p b 0.05). The average social interaction decreased with

Fig. 5. Relationships between average biomarker activity (EROD, GST, SOD and CAT) and two behavioral parameters (lethargy and socialinteraction) of juvenile seabream exposed to phenanthrene (0.11, 0.28 and 0.56 μM). Data were re-scaled by subtracting the control mean from eachmeasurement, and only the overall means are shown, with ● from first and ○ from the second study.

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increasing exposures up to 14% at 0.56 μM PHE, how-ever not statistically significant. This reduction in thesocial patterns of fish behaviors is probably because of thehigh incidence of lethargic fish. Aquaria with fish of high

individual apathy were always also the aquaria withlowest social interactions, i.e. biting, avoiding or chasingbehavior were reduced. A well-defined aggregative be-havior (Begout and Lagardere, 1995) is typical for the test

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species, and communities are often trying to establish aclear hierarchical structure of dominant and un-dominantgroup members.

As the number of fish with non-locomotor activityincreasedwith increasing PHE exposures, it is obvious thatthe number of active fish was reduced (Fig. 4C and D).Both swimming and rising were significantly lower at0.56 μM PHE, and data from both studies showed anexcellent agreement. Thus with increasing PHE concen-trations the active fish preferentially exhibited horizontalmovements. The data about the preferred position alsoindicated this: In the second study it was clearly themiddleof the aquarium.

3.4. Subcellular vs. behavioral responses

PHE caused effects both at biochemical and atindividual behavior level. Therefore, it is consequent torelate data from both levels for fixed concentrations, inorder to find out typical quantitative interrelationshipsbetween both. Indeed, despite the low treatment num-bers it was possible to identify trends between some ofthe enzymatic activities and the behavioral change inlethargy and social interactions (Fig. 5). All are based ondata values re-scaled to the control mean, with the zeroorigin of both axes corresponding to the control means.To enable a better visualization, only the means for eachPHE exposure are pictured, connected by a solid lineaccording to increasing PHE exposures (for 0.11 μM theaverage was chosen).

An increased EROD activity was related with ahigher lethargy in the fish samples only at low PHEconcentrations but markedly lowered for higher expo-sures (Fig. 5A). The opposite trend was observed forGST: Low PHE concentrations slightly increased thelethargic activity, but minimized the enzymatic activity,and the highest concentration (0.56 μM) produced thehighest number of lethargic fish and the highest GSTactivity (Fig. 5C). For SOD and CAT, we observedsimilar positive trends for changes in lethargy, withincreasing PHE concentrations causing enhanced enzy-matic activities (Fig. 5E and G). Social interaction wasreduced for the two highest PHE concentrations. Whenrelated to EROD activity (Fig. 5B), only the highestexposure showed a reduced enzymatic activity, whereatfor GST the reverse was observed: Low PHE exposurescaused no or only a slightly reduced enzymatic activity,and we detected only at 0.56 μM PHE an increased GSTactivity (Fig. 5D). For SOD and CAT, we observedsimilar negative trends for social interaction, with in-creasing PHE exposures provoking higher enzymaticactivities (Fig. 5F and H).

4. Discussion

In the field, fish can absorb PAHs fromwater via bodysurface or gills, from contaminated sediment and food. IfPAHs are taken up via gills, they are transported to theliver through the bloodstream, converted to water-soluble polar metabolites and excreted in the bile. Themetabolism of PAHs in fish might affect many sub-cellular processes and even influence biological levels ofhigh-order (reviewed by Van der Oost et al., 2003). Thismulti-response study investigated the biological impactsof PHE on subcellular and individual levels in fish.

4.1. Subcellular responses

The results of our experiments support a role forCYP1A (EROD) metabolism in the excretion and tox-icity of PHE (e.g. Hawkins et al., 2002; Shallaja andD'Silva, 2003; Oliveira et al., 2007). The concentration-dependent increase of PHE-type metabolites in bilefollowed by the increase of EROD activity at 0.11 μMindicates that PHE is metabolized in the liver ofseabream. Studies with similar waterborne PHE expo-sures to rainbow trout have evidenced that PHE is readilymetabolized by EROD and excreted in the bile (Hawkinset al., 2002), and Sun et al. (2006) demonstrated a shorthalf-time presence of this compound in whole-body ofCarassius auratus. However, in rainbow trout themetabolism of PHE was much more elevated when β-naphthoflavone (BNF) was used as co-exposure to PHE(Pangrekar et al., 2003; Billiard et al., 2004). Thissuggests that PHE is not such a strong cytochrome P450inducer as some of the commonly used model PAHsinducers (benzo[a]pyrene, BNF). Moreover, the degreeof stereoselectivity in the metabolism of PHE to benzo-ring dihydrodiols suggests that this compound, unlikebenzo[a]pyrene and chrysene, is metabolized by morethan one cytochrome P450 isoenzyme, presumably withdifferent stereoselectivities (Pangrekar et al., 2003). Thismight explain that some studies could not detect anycorrelation between crude oil contamination (e.g.naphthalene, PHE) and fish EROD activity. However,metabolites in bile correlate well with real exposures,demonstrating that the excretion of metabolites in bileis a suitable endpoint for oil contamination (Lee andAnderson, 2005). Our results strengthen these findingsbecause of the clear concordance between PHE exposureand biliary metabolites, but not for the relationshipbetween exposures and liver EROD activity.

The inactivation of biotransformation enzymes in fishliver at high PAH exposures (e.g. 0.9 μM BNF) is not anew finding (Haasch et al., 1993; Schlezinger and

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Stegeman, 2001; Gravato and Santos, 2002). Manyreasons might be responsible for such a phenomenon,e.g. either the co-occurrence of several CYP1A inducers(P450 gene subfamily) or the generation of metabolicproducts can interfere with the integrity of the enzymecausing its inactivation (see Stegeman and Hahn, 1994).On the other hand, the production of reactive oxygenspecies (ROS) linked to the CYP process (Schlezingeret al., 1999) caused an inactivation of scup CYP1A afterPCBs exposures (Schlezinger and Stegeman, 2001). Toour knowledge, this aspect was so far not sufficientlyconsidered for PAHs exposures, although it is well knownthat the PAH metabolism can be linked to the generationof ROS (Shi et al., 2005; Sun et al., 2006). For instance,they demonstrated that 0.3 μM PHE can induce 195% ofOH in the liver of C. auratus after 24 h. In addition, anincrease in ROS caused by PHE exposures could berelated to oxidative stress (Sun et al., 2006). Similarly, wedetected changes in oxidative stress enzymes in liver ofseabream, i.e. CAT activities increased in a dose-dependent manner suggesting an accumulation of H2O2.It is also likely that SODs are involved in the increasedlevels of such radicals, as these enzymes convertsuperoxide anions into H2O2 (Livingstone, 2001). Aswe have observed an increase of these antioxidantscavengers at high-PHE concentrations, we speculatethat the enhanced metabolism rate of PHE have createdpro-oxidant conditions which might have favored anEROD inactivation in livers. Although the highest testconcentration inhibited the EROD activity, neverthelesswe detected the highest levels of PHE in bile, revealingthat the chemical is still metabolized. The parallel increaseof liver GSTactivities in livers might be an indication thatGST is relevant for the phase II biotransformation of PHE.However, recently two other studies have analyzed theenzymatic activity of GST on PHE exposures in twospecies, tilapia (Shallaja and D'Silva, 2003) and oliveflounder (Jee and Kang, 2005), and found no relationshipbetween GST and PHE excretion. Although dihydrodiolderivates appear to be the most dominant metabolites ofPHE, often conjugated forms with sulphates andglucuronides (phase II conjugation) are detected in thebile of fish exposed to PAHs (see Watson et al., 2004).

GST is a multi-component enzyme, which is involvedin the detoxication of many xenobiotics (Van der Oostet al., 2003). For example, if the activity of this enzyme isincreased by 33% in high-PHE exposed tilapia, thensignificantly less liver damages are observed (Shallaja andD'Silva, 2003). Like for CAT and SOD, we observed inlivers of high-exposed seabream a markedly higherenzyme activity, probably in order to compensate theincrease of oxidative stress conditions because of higher

rates of PHEmetabolism. This is corroborated by findingsof Jee and Kang (2005), who detected increased levels ofGSTand CAT in olive flounder after two-week exposuresto PHE. The best-studied oxidative stress target is themembrane damaged through lipid peroxidation, which isinitiated by ROS that attack polyunsaturated fatty acids inmembranes and produce lipid breakdown products suchas MDA (Livingstone, 2001). In the current study, thelevels of MDA increased, even if non-significantly, tonearly 10% in low-exposed fish, but were similar to that ofthe controls when GST and the antioxidant enzymesdisplayed the maximal activity. Lipid peroxidation or theoxidation of polyunsaturated fatty acids is a veryimportant consequence of oxidative stress caused byhydrocarbon metabolism, e.g. Shi et al. (2005) demon-strated that lipid peroxidation is strongly related to ROSproduction when fish are exposed to naphthalene.

4.2. Behavioral responses and their relation tosubcellular activities

Animals are always behaving (Lehner, 1996), andthis might be the result of external and internal stimuli inorder to maintain their internal homeostasis. An externalstimulus, mainly associated with the presence of PHE inthe water, produced in our study not only severalbiochemical alterations in seabream, but also behavioralchanges. The most pronounced change in the individualmovement was the high increase of lethargic fish (up to39%) and, as a consequence thereof, it decreasedswimming activity with increasing PHE exposures.These changes in the individual behaviors had influ-enced negatively the social behavior: The more fish inthe group were lethargic, the less the social interactionsoccurred. These findings suggest that PHE exposurescan strongly influence the performance of normalseabream behavior. Indeed, it is not very surprisingthat toxicant stressors that might affect importantphysiological processes, e.g. neuronal, hormonal andmetabolic disruption, can interfere with the individualbehavior. Only a few empirical studies have beenconducted to evaluate the impact of PAH exposures onfish: e.g., anthracene and fluoranthene affect therespiration and osmoregulation in the gills (e.g. Barnettand Toews, 1978; Farr et al., 1995) and can lead to fishhyperactivity (Hall and Oris, 1991, Walker et al., 1998).Exposures of diluted hydrocarbon effluents and PAHs-spiked sediments caused also a reduced fish activity(Westlake et al., 1983; Hinkle-Conn et al., 1998). Thesefindings are in well agreement to our results and openthe discussion on how metabolism of PHE can be linkedto fish behavior.

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Alterations in the swimming activity are very oftenthe result of intrinsic changes in the fish metabolism,which necessitates a reduction of energy-costly move-ments (Sorensen et al., 1997). The detoxication of xeno-biotics is a process that requires elevated levels ofmetabolic resources, and, in order to respond to theseadditional metabolic requirements, it leads to an in-creased carbohydrate and protein metabolism (Scott andSloman, 2004). The behavior of an animal followsspecific physiological sequences, and active animalsprobably might compensate the costs of long exposuresby reducing their swimming activity. For instance,metals can interfere with the carbohydrate metabolism,and the energetic requirements for metal detoxicationresult into changing swimming activities in fish(Sorensen et al., 1997; Handy et al., 1999).

In ecotoxicological studies it is commonly reportedthat “trade-offs” between the metabolic costs ofchemical detoxication and other processes are vital tothe survival of the organism, such as respiration,growth and reproduction (Handy et al., 1999). Theobserved multi-level responses to PHE exposuresprovided evidences for a potential “trade-off” betweenPHE detoxication, via the biliary–hepatic system, andthe fish activity. Fig. 5 shows that after four-dayexposures the observed dose-response changes inseabream behaviors matched well with the measuredbiochemical events in liver and their mechanisticunderstanding: Low-exposed fish that behaved moresimilar to the controls had an increased EROD ac-tivity in the liver, but only low activities of GST, CATand SOD. This might be an indication that a lowmetabolism rate in the liver had only a minor impact tothe overall enzymatic processes, and thus affected notsignificantly the overall behavior of the fish. This issomewhat surprising as for 0.11 μM PHE already a 20-fold increased accumulation of biliary PHE-typemetabolites was measured. Only the much higheraccumulation of metabolites (up to two orders ofmagnitude) measured at higher exposures and thesupposed to be higher rates in the metabolism of PHEseem to have provoked marked changes in thebehaviors of fish (lethargy increased up to 38.6% andsocial interactivity decreased up to 13.9%). However,compared to the detected values at the lowest testconcentration (0.11 μM PHE), the EROD activity wasreduced at 0.28 μM PHE and for the highest exposureeven significantly lower than for the controls. Thisprovides evidence that EROD can tackle only withtoxicant stressors up to a certain size, but for moresevere exposures, other subcellular processes in liverare induced (see discussion above).

Hawkins et al. (2002) found that the inhibition ofendogenous EROD activity had biological consequencesby elicited signals of lethargy and loss of equilibrium inrainbow trout. The authors presumed that the parentcompound rather than the metabolic products of PHEbiotransformation was responsible for the observedbehavioral changes. Contrarily, in our study the inhibitionof EROD activity in liver of seabreamwas related with anincrease of PHE metabolites in the bile. Although it hasbeen shown that these metabolites can cause damages toliver cells (Shallaja and D'Silva, 2003), the increases ofGST, CAT and SOD activities at high exposures indicatethat the total detoxication in liver was probably enhanced.However, these added enzymatic activations might haverequired much more energetic costs in order to cope withthe chemical detoxication. The increased liver somaticindex (up to 20%, Table 1) is a clear indication of anabnormal high-metabolic activity in those exposed livers.Furthermore, Oliveira et al. (2007) exposed golden greymullet to 0.1 and 0.9μMPHE and observed an increase ofglucose plasma levels and liver EROD activities, whichsupports our findings that the overall metabolism of fishmight be affected when strong detoxication processes ofPHE are activated. The absence of the food during thefour-day exposures and as a consequence thereofmalnutrition of the animals at the end of the study wereprobably also responsible for less swimming activities, asclearly indicated in the second study by general reducedactivities both in controls and in treatments (Table 2).However, this circumstance affected all fish in the sameway and could thus not confound the observed concen-tration–response related changes.

The overall changes in behaviors might not benecessarily an indication for a reduced fitness, at leastwhen the animals are monitored only over a short time.However, in a long term, such responses might implynegative fails. For instance, lethargic juvenile have a lessoptimal feeding efficiency and thus reduced growth rateparameters (Purdy, 1989; Gregg et al., 1997; Hinkle-Connet al., 1998). We observed for the exposed seabream alower frequency of social interactions together with apreferred position in the middle of the aquarium, whichmight be a signal of a breakdown in the hierarchy structure(Sloman et al., 2005).

5. Conclusion

Thewhole pattern of biomarker responses gave insightin the fish and how their metabolism responded to short-term waterborne exposures of PHE. It seems that thischemical is readily metabolized in seabream liver throughthe EROD and GST biotransformation enzymes. The

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levels of PHE-type metabolites in the exposed bile are agood marker of PHE metabolism in the liver. A strongenhanced metabolism in the liver implies a reduction ofERODactivity, but an additional activation ofGST (phaseII enzyme), CAT and SOD enzymes. These biomarkerresponses walk along with severe behavioral changes infish. Thus, changes in the behavioral performance of fishseem to be the consequence of high-metabolic energeticcosts, which are inherent to detoxication processes ofPHE. However, the energetic demands for detoxicationcan reduce the energy for stores, growth and reproductionin long-term exposures (Berntssen et al., 2003). The effectpatterns observed in this study reflect real field situationsonly partly as it is more likely that fish are exposed to onlyvery low exposures of PHE. However, this compound isbarely the only potential chemical stressor in theenvironment, but often present in a complex mixture ofaromatic compounds (Zhang et al., 2004). Thus thequestion arises as to how the individual fish and the entirepopulation, respectively, can tackle with the joint effect ofmultiple PAHs, especially in case of chronic exposureconditions. In summary, changes in the non-locomotoractivity of fish revealed to be a non-evasive and sensitivebehavioral endpoint to PHE exposures. The sole use of abiomarker involves the danger to overlook relevanttoxicant responses, e.g. EROD activity was similar tothe controls atmediumPHEexposures, despite changes inthe individual behavior of fish. An integrated use ofbiomarkers can provide sufficient information that helpsus to understand the effects of PAHs on individualorganism and the population fitness.

Acknowledgements

This research presented here was sponsored by theproject POCI /MAR/56964/2004, co-financed byFEDER through Programa Operational Ciência eInovação 2010, fellowships SFRH/BPD/ 14419/2003and IEFP no. 013009. [SS]

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