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

Journal of Experimental Marine Biology and Ecology 347 (2007) 109 – 122 www.elsevier.com/locate/jembe

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-off” between detoxication 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

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consequence of higher energetic demands for the PHE detoxication. This illustrates how an integrated use of biomarkers can contribute 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 originate mainly from the production of synthetic chemicals and the use of fossil-energy. Of particular interest are polycyclic aromatic compounds (PAHs), a group of over 100 different chemicals that are formed during the incomplete burning of coal, oil and gas, garbage, or other organic substances like tobacco or charbroiled meat. In 2004, their quantities were estimated to exceed 2 million pounds in the US (US EPA, 2006). Organic xenobiotics are a potential threat to humans and the environment, especially with respect to PAHs which are suspected to be carcinogens (Albers, 2003). Because of their ability to absorb easily to organic materials (Law and Biscaya, 1994) they are commonly found as pollutants in soils, estuarine waters and sediments, and other terrestrial and aquatic sites. Most toxicological studies with PAHs have been investigated at molecular and cellular levels. For instance, enzymes that are part of the [Ah]-gene battery (e.g. CYP1A, enzymes of phase II conjugates, antioxidant enzymes) are often used as molecular biomarkers in order to investigate the influence of PAHs on the biochemical pathways and enzyme functioning in fish (reviewed by Whyte et al., 2000), and many studies have demonstrated that fish possess a well-developed MFO system that might efficiently detoxify a large number of xenobiotics, including PAHs. During the processes of detoxication, reactive metabolites can be produced and elicit toxicity through the generation of reactive oxygen species (ROS) and/or for binding covalently to cellular macromolecules such as DNA, RNA and protein (reviewed by Van der Oost et al., 2003). Biomarkers at molecular level are meaningful because they are able to respond quickly and often highly specific to chemical stressors (Van der Oost et al., 2003). However, their value is limited when we want to assess the impact of exposures for the whole organism, mainly as the link between biochemical responses and higher integral endpoints such as physiology, reproduction and behavior are too often unclear (Jensen et al., 1997; Livingstone, 2001). Studies looking on how the different levels of biological organization are related to each other thus improve the mechanistic understanding of toxicity

and their ecological consequences (Weis et al., 2001). Although it has been shown that PAHs can interfere on higher levels of organization in fish (e.g. Farr et al., 1995; Monteiro et al., 2000; Jee et al., 2004), the majority of the studies have investigated the impact of these compounds only at very specific levels of organization. Individual behavior is an integral response parameter that is linked to activities at biochemical levels, e.g. changes in the swimming activity of fish are commonly the result of damages in the nervous and hormonal control system, induced by metals and polychlorinated biphenyls—PCBs (Jensen et al., 1997, Weis et al., 2001). Although PAHs can affect the behavior performance of fish (Westlake et al., 1983; Farr et al., 1995; Hinkle-Conn et al., 1998), it remains unknown how this can be linked to disruptions at biochemical level. Typically, reproduction endpoints are used to assess chemical effects on population and community levels, but individual changes in behavior can provide similar information (Weis et al., 2001), as toxicants can disturb behavioral patterns that are essential for the fitness and survival of the entire population (Scott and Sloman, 2004). Thus, behavioral endpoints and their mechanistic understanding are an important step to analyze the connections between subtle biochemical changes in the organism and their ecological consequences. Phenanthrene (PHE) is a priority PAH, and, although not mutagenic or carcinogenic, it has been shown to be toxic to marine diatoms, gastropods, mussels, crustaceans, and fish (Albers, 2003; US EPA, 2006). Since PHE is the smallest tricyclic aromatic hydrocarbon to have a “bay-region” and a “K-region” (Ouyang, 2006), i.e. highly reactive regions of PAH molecules where the main carcinogenic species can be formed, it is commonly used as a model substrate for studies on metabolism of carcinogenic PAHs. We used PHE as a model compound in order to investigate the relationship between specific molecular processes (induction of biotransformation enzymes and oxidative stress) and the behavioral performance of fish at the individual level. Juvenile seabream (Sparus aurata) was used as model species because of its ability to yield reproducible behavior data under controlled conditions (Begout and Lagardere, 1995). Seabream is widely cultured in Europe. We performed two concentration-

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effect studies with PHE in a semi-static system, at which fish were exposed daily for 4 days. At the end, liver activities of ethoxyresorufin-O-deethlylase (EROD), total glutathione-S-transferase (GST) and PHE-type metabolites in bile were analyzed in order to gain insight into the detoxication and excretion mechanisms for PHE. Furthermore, we measured the responses of catalase (CAT) and superoxide dismutase (SOD) as indicators of oxidative stress and levels of lipid peroxidation (LP) as indication of damage. The physiological status of the fish was determined by the liver somatic index (LSI). We examined the behavior by recording visually the activity and social interactions of groups of fish per aquarium. 2. Material and methods 2.1. Test organisms and chemicals Gilthead seabream juveniles, S. aurata, L., were supplied from a commercial fish farm (TIMAR Lda., Setúbal, Portugal), where they had been raised till the weight of 1.0 g. All fish were from the same batch, and before dosing, they were kept under laboratory conditions in 60-l aquaria (density 2–3 g m− 3) supplied with filtered seawater (35 ± 2 ppm). The fish were fed daily with a maintenance ration of 2–3% body weight, and their average body weight during the exposures were 2.0 ± 0.2 g (first study, n = 75) and 2.2 ± 0.2 g (second study, n = 75). PHE (≥ 97% purity) was purchased from Aldrich (Milwankee, WI). All other chemicals were 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 glass aquaria at 16 ± 1 °C in filtered seawater (35 ± 2 ppm) under a photoperiod of 12 h light: 12 h dark. The aquaria were kept at semi-obscurity during the light periods in order to avoid PHE phototoxicity. Dissolved oxygen saturation (N 80%) and total ammonia concentrations (b 0.5 mg l− 1) were monitored weekly. Aquaria were constructed of glass, and the contact of other materials (e.g. silicon rubber tubing) with the test solutions was minimized. PHE was initially dissolved in acetone, and the stock solution was kept at −20 °C until prepared for the final exposure solutions in seawater. Exposures were daily renewed along with seawater (50% of total volume), and the solvent concentrations never exceeded 0.0014% in the aquaria. Water disposal from the aquaria was filtered through activated carbon before being delivered into the municipal sewage system.

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Before exposure, animals were acclimatized at the same conditions described for waterborne exposures for 24 h in 20-l aquaria and then five randomly chosen fish were placed in each test aquarium for 24 h with aeration (pre-exposure phase). Afterwards, fish were exposed to PHE for 4 days (post-exposure) (ASTM, 2003). Food was not provided during the acclimation and in the course of exposures. Aeration was provided with plastic tips placed 2 cm above the aquaria bottom. We conducted two consecutive studies within 1 month, at nominal concentrations of 0.11 and 0.56 μM PHE in the first study (0.02 and 0.1 mg l− 1, respectively), and 0.11 and 0.28 μM in the second (0.02 and 0.05 mg l− 1, respectively). Five fish per aquarium were used, and in 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). After 4 days of exposure, the animals were sacrificed for the subcellular analyses. 2.3. Sample preparation For the biochemical analysis, we always pooled liver tissues (50–100 mg wet weight) from two fish. Livers were homogenized in ten volumes of phosphate buffer (100 mM, pH 7.5) containing 1 mM EDTA, and afterwards centrifuged at 10,000 ×g for 20 min at 4 °C. We distributed the resulting postmitochondrial supernatants (PMS) into aliquots and stored them at − 80 °C prior to analysis. We used a sample volume of 250 μl for the lipid peroxidation assay and of 40 μl for the enzymatic measurements (EROD, GST, CAT, and SOD). Total protein concentration was determined in PMS supernatants according to the Lowry method (Lowry et al., 1951) and adapted to microplates using bovine serum albumin as standard. We collected bile samples by incising the gall bladder and stored them at − 80 °C until the analysis. We excised and weighed individual livers and determined the liver somatic index (LSI) as the percentage ratio of liver weight to body weight. 2.4. Biochemical assays EROD activity was measured by the fluorimetric method described in Solé et al. (2000). PMS liver samples (25 μl) were incubated at 30 °C 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 μM 7ethoxyresorufin. The reaction was stopped by adding 1 ml of ice-cold acetone, samples were centrifuged at 400 ×g and 7-hydroethoxyresorufin fluorescence was

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determined at 530/585 nm excitation/emission wavelengths. 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 to the method of Habig et al. (1974) and adapted to microplates (Frasco and Guilhermino, 2002). 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, we added 0.2 ml of the reaction mixture to 0.1 ml of the sample, with a final concentration of 1 mM GSH and 1 mM CDNB in the assay. GST was measured using CDNB as change in OD/min at 340 nm (ε = 9.6 mM− 1 cm− 1) and expressed as nmol min− 1 mg prot− 1. Catalase (CAT) activity was measured by the decrease in absorbance at 240 nm because of H2O2 consumption (ε = 40 M− 1 cm− 1). For the reaction, we used 67.5 mM potassium phosphate buffer (pH 7.5) and 12.5 mM H2O2, and initiated it with the addition of the sample. We expressed the CAT as μmol min− 1 mg prot− 1. SOD activity was determined in the PMS liver fraction as inhibition of cytochrome c reduction at 550 nm (McCord and Fridovich, 1969), adopted to microplate (Ferreira et al., 2005). The reaction contained phosphate buffer (50 mM, pH 7.8), 50 μM hypoxanthine, 1.98 mU ml− 1 xanthine oxidase and 10 μM cytochrome c. We measured the relative activity in units of SOD (U mg prot− 1), with one unit SOD being the amount of sample causing a 50% inhibition of cytochrome c reduction under the standard conditions of the assay. Tissue lipid peroxides (malondialdehyde—MDA equivalents) were measured in PMS by the thiobarbituric acid method (Niki, 2000). Subsamples of tissue homogenate were incubated with 100% TCA, and after centrifugation the supernatant was incubated for 30 min at 100 °C with 1% TBA, 0.05 M NaOH and 0.025% BHT. The supernatant (organic layer) was taken and its absorbance measured at 532 nm (ε = 1.54 × 105 M− 1 cm− 1, Halliwell and Gutteridge, 1999). We expressed the lipid peroxidation (LP) as MDA equivalents per mg liver (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 to 1:100,000. 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 and the same person recorded their behavior in all studies. In order to avoid a recording bias, exposures were unknown to this person. Individual fish responses were monitored each day for 2 h from 10 to 12 a.m. by recording visually at every 12 min their behavioral activity and spatial distribution as an “all-or-none” response, obtaining 10 counts for each aquarium per session. After recording, aqueous PHE and solvent control solutions were administered to the aquaria. The study was completed after 5 days, with one pre- and four post-exposure data sets of behavioral records for each aquarium. Check sheets were used to record behavioral observations. The basic design of the check sheet was a grid, with columns denoting successive sample intervals and rows denoting the behavior endpoints defined (Martin and Bateson, 1993). Each individual behavioral activity was categorized into three types: (i) swimming patterns, defined in terms of horizontal movements (swimming) and vertical movements (rising), (ii) lethargy, described as a non-locomotory activity by the absence of detectable body movements, and (iii) social patterns (social interaction), i.e. avoiding, biting or chasing behavior. Such behavioral categories can be affected by environmental contaminants, including PAHs (Sorensen et al., 1997; Sloman et al., 2003). Additionally, the position of each fish in the aquaria was recorded (bottom, middle, surface and near the aeration filter) (Yilmaz et al., 2004). 2.7. Statistical analyses Biomarker data were examined for normal distribution and homogeneity of variance (Shapiro–Wilk's and Bartlett's tests), and if required, data were log transformed. Dunnett's test (α = 5%) was then employed to determine whether any of the treatment groups differed in relation to the solvent controls, with aquarium always a nested factor in data analysis. We proved that the behavioral count data follow an overdispersed binomial distribution, which lead to difficulties for the data analysis: As the experimental design is nested, and differences between control and exposure means are of main interest, it formulates an unsolved problem in statistics and ruled out a powerful data analysis. Instead, we used the aquaria means as statistical units and assessed the differences between controls and exposures by the nonparametric Kruskal– Wallis test. A further quantitative difficulty was the correct choice of the control reference for the post-

15

25

6–10

10

9

10

25

7–10

9–10

8

10

13–15

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

One-sided

One-sided Phenanthrene-type metabolites (μg ml− 1, ppm)

Two-sided PL (nmol MDA g− 1)

LSI (%)

One-sided

One-sided CAT (µmol min− 1 prot− 1)

SOD (U mg prot− 1)

Two-sided GST (nmol min− 1 prot− 1)

⁎ Significant at α = 5%.

Replicates 0.28 μM

52.64 [37.33;74.24] 133.9 [117.7;152.4] 49.2 ⁎ [47.2;51.2] 16.97 [14.16;20.33] 62.9 [52.4;75.6] 1.38 [1.32;1.45] 8928 ⁎ [7591;10501] 72.25 ⁎ [48.65;107.30] 124.8 [108.3;143.8] 51.9 ⁎ [48.2;55.9] 16.77 [14.87;18.92] 65.2 [42.8;99.3] 1.18 [1.10;1.28] 3389 [2704;4248] 46.85 [38.29;57.33] 140.7 [107.5;184.2] 42.0 [39.2;45.0] 15.64 [13.88;17.63] 49.7 [41.7;59.3] 1.21 [1.11;1.33] 151.0 [110.7;206.1]

0.11 μM Control Replicates

8–10

3. Results

42.6 ⁎ [29.9;60.7] 160.2 [138.1;185.9] 57.9 ⁎ [52.2;64.3] 17.9 [15.1;21.2] 58.0 [37.8;88.9] 1.61 ⁎ [1.44;1.80] 10252 ⁎ [8324;12625]

0.56 μM 0.11 μM

exposed fish, with typically two possible approaches. One approach is to compute the difference of response values for each fish group in a given aquarium (treatments and controls) with its own control values prior to exposure, and a mean difference is then estimated for a given behavioral response variable in each aquarium at a given time (individual pre–post comparison). This mean difference is then compared to the mean difference observed in the control fish aquarium (ASTM, 2003). This approach assumes that a fish (or group) “remembers” its preexposure behavior. Alternatively, in the other approach, the post-exposure data for the treatments are compared with the post-control data (post–post comparison), which requires no pre-exposure information. Because of the long exposure duration and the absence of food for the juvenile fish, we found the latter approach also suitable and used therefore both approaches. All analyses were performed using the SAS procedure PROC GENMOD and PROC GLM (SAS version 9, SAS Institute Inc, Cary, NC, USA).

104.0 [79.7;135.7] 107.9 [87.6;132.7] 51.5 [44.9;59.1] 15.41 [14.79;16.04] 64.4 [53.3;77.9] 1.46 [1.29;1.65] 2728 ⁎ [2254;3301] Two-sided EROD (pmol min− 1 prot− 1)

Control

113

71.1 [59.9;84.2] 143.5 [113.3;181.7] 45.2 [37.0;55.3] 15.4 [14.0;16.9] 58.3 [49.0;69.3] 1.34 [1.25;1.44] 102.5 [87.5;120.0]

Second study First study Dunnett Endpoint

Table 1 Mean effects for phenanthrene exposures to juvenile seabream (always nested analysis with factor aquarium, data always log transformed)

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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 seabream exposed to phenanthrene (0.11, 0.28 and 0.56 μM). Data were rescaled by subtracting the control mean from each measurement. Error bars show the mean with 95% confidence belts, with ● from first and ○ from the second study.

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0.56 μM). Both studies were conducted successfully, and we observed neither fish mortality nor any evidences for fish infections or other diseases. Statistical results about the average enzymatic activities in juvenile fish are given in Table 1, together with the 95% confidence intervals, the number of replicates and all statistical test decisions. In order to achieve better data comparability between both studies, we re-scaled these data by subtracting the control mean from the data observations, and results are shown in Figs. 1 and 2. Outcomes from the analysis of PHE-type metabolites in fish bile are pictured in Fig. 3. The behavioral activities of fish before and after 4 days of dosing are summarized in Table 2, and the corresponding relative changes to the controls are presented in Table 3 and Fig. 4. In Fig. 5, the

Fig. 3. Concentration–response data of phenanthrene-type metabolites (equivalents, μg ml− 1) in bile of juvenile seabream exposed to phenanthrene (0.11, 0.28 and 0.56 μM). The black dots represent the means, connected by a smoothing solid line.

relationships between observed behavioral changes and measured enzyme activities are pictured exemplarily for eight selected cases, i.e. two behavioral parameters (lethargy and social interactions) are related to EROD, GST, CAT and SOD responses. 3.1. Subcellular responses

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

Pooled data for EROD activity in seabream showed a non-monotonic concentration–response relationship (Fig. 1A), as for the lowest tested concentration (0.11 μM PHE) the measurements were about 1.5-fold higher than control values, while at the highest tested concentration (0.56 μM PHE) the activity was significantly suppressed ( p b 0.05). Although we detected the observed increase as statistically significant only for data from the second study, both studies yielded similar mean values. Indeed a pooled data analysis for the rescaled data (with study as co-factor in the model) confirmed the statistical significance (data not shown). Concentration–response data for GST activity indicate the opposite pattern: The measurements were in average 25% lower for 0.11 μM PHE, but 14% higher for 0.56 μM PHE (Fig. 1B). However, because of higher data variation the statistical power was not sufficiently high to detect these small differences as statistically significant. Both antioxidant enzymes, CAT and SOD, increased with higher PHE concentrations (Fig. 2A and B), with CAT levels in liver at the highest tested concentration around 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 levels were measured.

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

Second study 0.11 μM

Control

0.56 μM

0.11 μM

Control

0.28 μM

Pre

Post

Pre

Post

Pre

Post

Pre

Post

Pre

Post

Pre

Post

24.8 6.8 34.9 33.5

29.6 9.8 29.6 30.9

30.4 10.8 34.0 24.8

23.7 7.0 38.3 30.8

20.4 8.0 41.6 30.0

12.3 2.4 68.2 17.0

46.8 26.0 12.4 14.8

41.2 11.2 19.2 28.4

49.6 13.6 17.2 19.6

34.8 6.4 23.6 35.2

46.0 27.6 12.8 13.6

28.4 6.4 42.4 22.8

Position in the aquaria Surface 0.0 Middle 88.2 Bottom 11.8 Aeration filter 0.0

2.4 88.1 7.1 2.4

0.0 77.6 21.6 0.8

1.2 82.3 11.4 5.1

1.6 76.4 21.6 0.4

0.0 87.2 9.2 3.6

0.8 88.4 10.8 0.0

2.8 82.8 14.4 0.0

0.0 93.6 6.4 0.0

2.0 95.2 2.8 0.0

0.4 94.8 4.8 0.0

1.2 92.8 6.0 0.0

Behavioral activities Swimming Rising Lethargy Social interactions

a b

Fish were recorded before dosing. Fish were recorded after four-day exposures.

Levels of liver lipid peroxide, measured in terms of MDA, indicate a non-monotonic concentration–response pattern similarly to that of EROD activity (Fig. 2C): In both studies, we measured highest levels of MDA at 0.11 μM PHE, but observed less pronounced levels for higher test concentrations. However, the increase was at maximum only 8–10% above controls, and, with a coefficient variation of 20–30% in the controls, thus far below the minimal detection limit for statistics. The analysis of fluorescent aromatic compounds (FACs) showed clearly the presence of PHE-type metabolites in fish bile (Fig. 3), with average levels of 0.1–0.15 μg ml− 1 in the controls. Moreover, a clear concentration-dependent accumulation was evident, and at 0.11 μM PHE already a 26-fold higher level in bile fluid was measured (2.7 μg ml− 1, p b 0.05). The Table 3 Behavioural changes in juvenile seabream after four-day exposures to phenanthrene Control First study

Second study

0.11 μM 0.56 μM 0.11 μM 0.28 μM Behavioral activities Swimming Rising Lethargy Social interactions

0 0 0 0

Position in the aquaria Surface 0 Middle 0 Bottom 0 Aeration filter 0

− 5.9 − 2.8 8.7 ⁎ − 0.1 − 1.2 − 5.8 4.3 2.7

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

− 17.3 ⁎ − 7.4 ⁎ 38.6 ⁎ − 13.9

−6.4 −4.6 4.4 6.8

− 12.8 − 4.8 23.2 ⁎ − 5.6

− 2.4 − 0.9 2.1 1.2

−0.8 12.4 ⁎ − 11.6 0.0

− 1.6 10.0 − 8.4 0.0

concentration–response curve indicates that the accumulated levels in bile reached nearly a steady state at 0.28 μM PHE, with higher exposure concentrations producing only a minor increase in fluorescence. 3.2. Liver somatic index Compared to control fish, the PHE exposures produced only slight changes in liver somatic index (LSI), and only the highest concentration at 0.56 μM PHE caused a significant increase (Table 1). 3.3. Behavioral responses We have summarized the recorded behavioral activities for the tested juvenile fish in Table 2, for both studies, and always before and after exposures. Each value represents the mean percentage activity from a total of 25 fish, observed in five aquaria. Behavioral performance parameter are categorized according to the swimming activities (swimming, rising, lethargy, and social 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 the recording period in the middle of the aquarium (82–95%), and the fish avoided the surface and the proximity to aeration filter. However, the swimming activities differed between the studies: In the first study, around 70% of the control fish showed behaviors like lethargy and social interactions at study begin, which were reduced to 60% after 4 days. However, in the second study these activities were reduced to 27.2% prior dosing and to 47.6% after 4 days. The reason for these differences remains unclear.

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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 an aquarium, each black dot (●) the overall mean from the first study and grey dot ( ) the overall mean from the second study. For a better visuality, data from second study are shifted slightly to the right.

Furthermore, the differences in swimming activities before and after dosing were in the controls from the second study much more pronounced, probably because of the generally reduced non-locomotor activities in the second study. Although these differences between the studies clearly deny a simple data pooling, both studies nevertheless have produced very similar concentration– response pattern. This can be identified in the best way when not only the absolute concentration–response values are compared, but also changes to average preor post-control activity are considered. Table 3 shows the changes for post-exposure data when compared to the average control activity after 4 days, which are visualized for the swimming parameters in Fig. 4. It shows clearly that PHE has caused a change in the overall performance of the fish samples, as all behavior activities were affected in a dose-response way. The relatively good agreement between 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 the changes solely on a pre–post comparison (normalized to the mean difference of the controls) then these concentration–response relationships were masked by a huge data variation and changes are not anymore identifiable (data not shown). This indicates that the original behavioral pattern of fish in an individual aquarium was not maintained over the study duration, and consequently the pre-exposure information is not required for a meaningful concentration–response analysis. The observed concentration–response pattern varied substantially between the locomotor and non-locomotor activities. The clearest results were obtained for the increased 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

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

increasing exposures up to 14% at 0.56 μM PHE, however not statistically significant. This reduction in the social patterns of fish behaviors is probably because of the high incidence of lethargic fish. Aquaria with fish of high

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

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species, and communities are often trying to establish a clear hierarchical structure of dominant and un-dominant group members. As the number of fish with non-locomotor activity increased with increasing PHE exposures, it is obvious that the number of active fish was reduced (Fig. 4C and D). Both swimming and rising were significantly lower at 0.56 μM PHE, and data from both studies showed an excellent agreement. Thus with increasing PHE concentrations the active fish preferentially exhibited horizontal movements. The data about the preferred position also indicated this: In the second study it was clearly the middle of the aquarium.

4. Discussion In the field, fish can absorb PAHs from water via body surface or gills, from contaminated sediment and food. If PAHs are taken up via gills, they are transported to the liver through the bloodstream, converted to watersoluble polar metabolites and excreted in the bile. The metabolism of PAHs in fish might affect many subcellular processes and even influence biological levels of high-order (reviewed by Van der Oost et al., 2003). This multi-response study investigated the biological impacts of PHE on subcellular and individual levels in fish. 4.1. Subcellular responses

3.4. Subcellular vs. behavioral responses PHE caused effects both at biochemical and at individual behavior level. Therefore, it is consequent to relate data from both levels for fixed concentrations, in order to find out typical quantitative interrelationships between both. Indeed, despite the low treatment numbers it was possible to identify trends between some of the enzymatic activities and the behavioral change in lethargy and social interactions (Fig. 5). All are based on data values re-scaled to the control mean, with the zero origin of both axes corresponding to the control means. To enable a better visualization, only the means for each PHE exposure are pictured, connected by a solid line according to increasing PHE exposures (for 0.11 μM the average was chosen). An increased EROD activity was related with a higher lethargy in the fish samples only at low PHE concentrations but markedly lowered for higher exposures (Fig. 5A). The opposite trend was observed for GST: Low PHE concentrations slightly increased the lethargic activity, but minimized the enzymatic activity, and the highest concentration (0.56 μM) produced the highest number of lethargic fish and the highest GST activity (Fig. 5C). For SOD and CAT, we observed similar positive trends for changes in lethargy, with increasing PHE concentrations causing enhanced enzymatic activities (Fig. 5E and G). Social interaction was reduced for the two highest PHE concentrations. When related to EROD activity (Fig. 5B), only the highest exposure showed a reduced enzymatic activity, whereat for GST the reverse was observed: Low PHE exposures caused no or only a slightly reduced enzymatic activity, and we detected only at 0.56 μM PHE an increased GST activity (Fig. 5D). For SOD and CAT, we observed similar negative trends for social interaction, with increasing PHE exposures provoking higher enzymatic activities (Fig. 5F and H).

The results of our experiments support a role for CYP1A (EROD) metabolism in the excretion and toxicity of PHE (e.g. Hawkins et al., 2002; Shallaja and D'Silva, 2003; Oliveira et al., 2007). The concentrationdependent increase of PHE-type metabolites in bile followed by the increase of EROD activity at 0.11 μM indicates that PHE is metabolized in the liver of seabream. Studies with similar waterborne PHE exposures to rainbow trout have evidenced that PHE is readily metabolized by EROD and excreted in the bile (Hawkins et al., 2002), and Sun et al. (2006) demonstrated a short half-time presence of this compound in whole-body of Carassius auratus. However, in rainbow trout the metabolism of PHE was much more elevated when βnaphthoflavone (BNF) was used as co-exposure to PHE (Pangrekar et al., 2003; Billiard et al., 2004). This suggests that PHE is not such a strong cytochrome P450 inducer as some of the commonly used model PAHs inducers (benzo[a]pyrene, BNF). Moreover, the degree of stereoselectivity in the metabolism of PHE to benzoring dihydrodiols suggests that this compound, unlike benzo[a]pyrene and chrysene, is metabolized by more than one cytochrome P450 isoenzyme, presumably with different stereoselectivities (Pangrekar et al., 2003). This might explain that some studies could not detect any correlation 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 bile is a suitable endpoint for oil contamination (Lee and Anderson, 2005). Our results strengthen these findings because of the clear concordance between PHE exposure and biliary metabolites, but not for the relationship between exposures and liver EROD activity. The inactivation of biotransformation enzymes in fish liver at high PAH exposures (e.g. 0.9 μM BNF) is not a new finding (Haasch et al., 1993; Schlezinger and

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Stegeman, 2001; Gravato and Santos, 2002). Many reasons might be responsible for such a phenomenon, e.g. either the co-occurrence of several CYP1A inducers (P450 gene subfamily) or the generation of metabolic products can interfere with the integrity of the enzyme causing its inactivation (see Stegeman and Hahn, 1994). On the other hand, the production of reactive oxygen species (ROS) linked to the CYP process (Schlezinger et al., 1999) caused an inactivation of scup CYP1A after PCBs exposures (Schlezinger and Stegeman, 2001). To our knowledge, this aspect was so far not sufficiently considered for PAHs exposures, although it is well known that the PAH metabolism can be linked to the generation of ROS (Shi et al., 2005; Sun et al., 2006). For instance, they demonstrated that 0.3 μM PHE can induce 195% of OH in the liver of C. auratus after 24 h. In addition, an increase in ROS caused by PHE exposures could be related to oxidative stress (Sun et al., 2006). Similarly, we detected changes in oxidative stress enzymes in liver of seabream, i.e. CAT activities increased in a dosedependent manner suggesting an accumulation of H2O2. It is also likely that SODs are involved in the increased levels of such radicals, as these enzymes convert superoxide anions into H2O2 (Livingstone, 2001). As we have observed an increase of these antioxidant scavengers at high-PHE concentrations, we speculate that the enhanced metabolism rate of PHE have created pro-oxidant conditions which might have favored an EROD inactivation in livers. Although the highest test concentration inhibited the EROD activity, nevertheless we detected the highest levels of PHE in bile, revealing that the chemical is still metabolized. The parallel increase of liver GST activities in livers might be an indication that GST is relevant for the phase II biotransformation of PHE. However, recently two other studies have analyzed the enzymatic activity of GST on PHE exposures in two species, tilapia (Shallaja and D'Silva, 2003) and olive flounder (Jee and Kang, 2005), and found no relationship between GST and PHE excretion. Although dihydrodiol derivates appear to be the most dominant metabolites of PHE, often conjugated forms with sulphates and glucuronides (phase II conjugation) are detected in the bile of fish exposed to PAHs (see Watson et al., 2004). GST is a multi-component enzyme, which is involved in the detoxication of many xenobiotics (Van der Oost et al., 2003). For example, if the activity of this enzyme is increased by 33% in high-PHE exposed tilapia, then significantly less liver damages are observed (Shallaja and D'Silva, 2003). Like for CAT and SOD, we observed in livers of high-exposed seabream a markedly higher enzyme activity, probably in order to compensate the increase of oxidative stress conditions because of higher

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rates of PHE metabolism. This is corroborated by findings of Jee and Kang (2005), who detected increased levels of GST and CAT in olive flounder after two-week exposures to PHE. The best-studied oxidative stress target is the membrane damaged through lipid peroxidation, which is initiated by ROS that attack polyunsaturated fatty acids in membranes and produce lipid breakdown products such as MDA (Livingstone, 2001). In the current study, the levels of MDA increased, even if non-significantly, to nearly 10% in low-exposed fish, but were similar to that of the controls when GST and the antioxidant enzymes displayed the maximal activity. Lipid peroxidation or the oxidation of polyunsaturated fatty acids is a very important consequence of oxidative stress caused by hydrocarbon metabolism, e.g. Shi et al. (2005) demonstrated that lipid peroxidation is strongly related to ROS production when fish are exposed to naphthalene. 4.2. Behavioral responses and their relation to subcellular activities Animals are always behaving (Lehner, 1996), and this might be the result of external and internal stimuli in order to maintain their internal homeostasis. An external stimulus, mainly associated with the presence of PHE in the water, produced in our study not only several biochemical alterations in seabream, but also behavioral changes. The most pronounced change in the individual movement was the high increase of lethargic fish (up to 39%) and, as a consequence thereof, it decreased swimming activity with increasing PHE exposures. These changes in the individual behaviors had influenced negatively the social behavior: The more fish in the group were lethargic, the less the social interactions occurred. These findings suggest that PHE exposures can strongly influence the performance of normal seabream behavior. Indeed, it is not very surprising that toxicant stressors that might affect important physiological processes, e.g. neuronal, hormonal and metabolic disruption, can interfere with the individual behavior. Only a few empirical studies have been conducted to evaluate the impact of PAH exposures on fish: e.g., anthracene and fluoranthene affect the respiration and osmoregulation in the gills (e.g. Barnett and Toews, 1978; Farr et al., 1995) and can lead to fish hyperactivity (Hall and Oris, 1991, Walker et al., 1998). Exposures of diluted hydrocarbon effluents and PAHsspiked sediments caused also a reduced fish activity (Westlake et al., 1983; Hinkle-Conn et al., 1998). These findings are in well agreement to our results and open the discussion on how metabolism of PHE can be linked to fish behavior.

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Alterations in the swimming activity are very often the result of intrinsic changes in the fish metabolism, which necessitates a reduction of energy-costly movements (Sorensen et al., 1997). The detoxication of xenobiotics is a process that requires elevated levels of metabolic resources, and, in order to respond to these additional metabolic requirements, it leads to an increased carbohydrate and protein metabolism (Scott and Sloman, 2004). The behavior of an animal follows specific physiological sequences, and active animals probably might compensate the costs of long exposures by reducing their swimming activity. For instance, metals can interfere with the carbohydrate metabolism, and the energetic requirements for metal detoxication result into changing swimming activities in fish (Sorensen et al., 1997; Handy et al., 1999). In ecotoxicological studies it is commonly reported that “trade-offs” between the metabolic costs of chemical detoxication and other processes are vital to the survival of the organism, such as respiration, growth and reproduction (Handy et al., 1999). The observed multi-level responses to PHE exposures provided evidences for a potential “trade-off” between PHE detoxication, via the biliary–hepatic system, and the fish activity. Fig. 5 shows that after four-day exposures the observed dose-response changes in seabream behaviors matched well with the measured biochemical events in liver and their mechanistic understanding: Low-exposed fish that behaved more similar to the controls had an increased EROD activity in the liver, but only low activities of GST, CAT and SOD. This might be an indication that a low metabolism rate in the liver had only a minor impact to the overall enzymatic processes, and thus affected not significantly the overall behavior of the fish. This is somewhat surprising as for 0.11 μM PHE already a 20fold increased accumulation of biliary PHE-type metabolites was measured. Only the much higher accumulation of metabolites (up to two orders of magnitude) measured at higher exposures and the supposed to be higher rates in the metabolism of PHE seem to have provoked marked changes in the behaviors of fish (lethargy increased up to 38.6% and social interactivity decreased up to 13.9%). However, compared to the detected values at the lowest test concentration (0.11 μM PHE), the EROD activity was reduced at 0.28 μM PHE and for the highest exposure even significantly lower than for the controls. This provides evidence that EROD can tackle only with toxicant stressors up to a certain size, but for more severe exposures, other subcellular processes in liver are induced (see discussion above).

Hawkins et al. (2002) found that the inhibition of endogenous EROD activity had biological consequences by elicited signals of lethargy and loss of equilibrium in rainbow trout. The authors presumed that the parent compound rather than the metabolic products of PHE biotransformation was responsible for the observed behavioral changes. Contrarily, in our study the inhibition of EROD activity in liver of seabream was related with an increase of PHE metabolites in the bile. Although it has been shown that these metabolites can cause damages to liver cells (Shallaja and D'Silva, 2003), the increases of GST, CAT and SOD activities at high exposures indicate that the total detoxication in liver was probably enhanced. However, these added enzymatic activations might have required much more energetic costs in order to cope with the chemical detoxication. The increased liver somatic index (up to 20%, Table 1) is a clear indication of an abnormal high-metabolic activity in those exposed livers. Furthermore, Oliveira et al. (2007) exposed golden grey mullet to 0.1 and 0.9 μM PHE and observed an increase of glucose plasma levels and liver EROD activities, which supports our findings that the overall metabolism of fish might be affected when strong detoxication processes of PHE are activated. The absence of the food during the four-day exposures and as a consequence thereof malnutrition of the animals at the end of the study were probably also responsible for less swimming activities, as clearly indicated in the second study by general reduced activities both in controls and in treatments (Table 2). However, this circumstance affected all fish in the same way and could thus not confound the observed concentration–response related changes. The overall changes in behaviors might not be necessarily an indication for a reduced fitness, at least when the animals are monitored only over a short time. However, in a long term, such responses might imply negative fails. For instance, lethargic juvenile have a less optimal feeding efficiency and thus reduced growth rate parameters (Purdy, 1989; Gregg et al., 1997; Hinkle-Conn et al., 1998). We observed for the exposed seabream a lower frequency of social interactions together with a preferred position in the middle of the aquarium, which might be a signal of a breakdown in the hierarchy structure (Sloman et al., 2005). 5. Conclusion The whole pattern of biomarker responses gave insight in the fish and how their metabolism responded to shortterm waterborne exposures of PHE. It seems that this chemical is readily metabolized in seabream liver through the EROD and GST biotransformation enzymes. The

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levels of PHE-type metabolites in the exposed bile are a good marker of PHE metabolism in the liver. A strong enhanced metabolism in the liver implies a reduction of EROD activity, but an additional activation of GST (phase II enzyme), CAT and SOD enzymes. These biomarker responses walk along with severe behavioral changes in fish. Thus, changes in the behavioral performance of fish seem to be the consequence of high-metabolic energetic costs, which are inherent to detoxication processes of PHE. However, the energetic demands for detoxication can reduce the energy for stores, growth and reproduction in long-term exposures (Berntssen et al., 2003). The effect patterns observed in this study reflect real field situations only partly as it is more likely that fish are exposed to only very low exposures of PHE. However, this compound is barely the only potential chemical stressor in the environment, but often present in a complex mixture of aromatic compounds (Zhang et al., 2004). Thus the question arises as to how the individual fish and the entire population, respectively, can tackle with the joint effect of multiple PAHs, especially in case of chronic exposure conditions. In summary, changes in the non-locomotor activity of fish revealed to be a non-evasive and sensitive behavioral endpoint to PHE exposures. The sole use of a biomarker involves the danger to overlook relevant toxicant responses, e.g. EROD activity was similar to the controls at medium PHE exposures, despite changes in the individual behavior of fish. An integrated use of biomarkers can provide sufficient information that helps us to understand the effects of PAHs on individual organism and the population fitness. Acknowledgements This research presented here was sponsored by the project POCI /MAR/56964/2004, co-financed by FEDER through Programa Operational Ciência e Inovação 2010, fellowships SFRH/BPD/ 14419/2003 and IEFP no. 013009. [SS] References Albers, P.H., 2003. Petroleum and individual polycyclic aromatic hydrocarbons. In: Hoffman, D.J., Rattner, B.A., Burton, G.A., Cairns, J. (Eds.), Handbook of Ecotoxicology. Lewis Publishers, New York, pp. 1–32. American Society for Testing and Materials (ASTM), 2003. Standard guide for ventilation behavioral toxicology testing of freshwater fish. E 1768-95 (reapproved 2003). Annual Book of ASTM Standards, Water and Environmental Technology, vol.11.05. American Society for Testing Materials, West Conshohocken, PA. Barnett, J., Toews, D., 1978. Effects of crude-oil and dispersant, Oilsperse 43, on respiration and coughing rates in Atlantic salmon (Salmo salar). Can. J. Zool. 56, 307–310.

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