The effects of carbamazepine on macroinvertebrate species: Comparing bivalves and polychaetes biochemical responses

Water Research 85 (2015) 137e147

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The effects of carbamazepine on macroinvertebrate species: Comparing bivalves and polychaetes biochemical responses ^ tia Velez a, Va ^nia Calisto b, Rosa Freitas a, *, Angela Almeida a, Adília Pires a, Ca Rudolf J. Schneider c, Valdemar I. Esteves b, Frederick J. Wrona a, d, Etelvina Figueira a, Amadeu M.V. M. Soares a a

Department of Biology & CESAM, University of Aveiro, Aveiro, Portugal Department of Chemistry & CESAM, University of Aveiro, Aveiro, Portugal BAM Federal Institute for Materials Research and Testing, Richard-Willstaetter -Str. 11, Berlin, Germany d Department of Geography, University of Victoria, 3800 Finnerty Road, David Turpin Building, Victoria, BC V8P 5C2, Canada b c

a r t i c l e i n f o

a b s t r a c t

Article history: Received 14 May 2015 Received in revised form 3 July 2015 Accepted 1 August 2015 Available online 5 August 2015

In the present study, the bivalve Scrobicularia plana and the polychaete Diopatra neapolitana were exposed to an increasing carbamazepine (CBZ) concentration gradient. Both species are among the most widely used bioindicators, and CBZ is one of the most commonly found drugs in the aquatic environment. After a chronic exposure (28 days), the results obtained revealed that CBZ induced biochemical alterations in both species. Our findings demonstrated that S. plana and D. neapolitana reduced the CBZ accumulation rate at higher CBZ concentrations, probably due to their capacity to decrease their feeding rates at stressful conditions. Nevertheless, this defence mechanism was not enough to prevent both species from oxidative stress. In fact, S. plana and D. neapolitana were not able to efficiently activate their antioxidant defence mechanisms which resulted in the increase of lipid peroxidation, especially at the highest CBZ concentrations. Comparing both species, it seems that S. plana was the most sensitive species since stronger biochemical alterations were observed in this species. © 2015 Elsevier Ltd. All rights reserved.

Keywords: Biomarkers Oxidative stress Scrobicularia plana Diopatra neapolitana Pharmaceutical drugs

1. Introduction Pharmaceutical drugs are an emerging class of environmental contaminants (Fent et al., 2006). Many of these compounds are not completely eliminated after their consumption by humans or animals (Boxall et al., 2004; Heberer, 2002; Heberer et al., 2002), and posteriorly excreted as slightly transformed product or even unchanged to wastewater treatment plants (WWTPs). Since conventional WWTPs are generally not effective in removing many pharmaceuticals and their metabolites, effluents resulting from these treatments are considered an important continuous source of drugs input into aquatic ecosystems (Brun et al., 2006). Although several studies have been published regarding the detection and quantification of these pharmaceuticals both in effluent and receiving waters (Brun et al., 2006; Fent et al., 2006;

* Corresponding author. Departamento de Biologia, Universidade de Aveiro, rio de Santiago, 3810-193 Aveiro, Portugal. Campus Universita E-mail address: [email protected] (R. Freitas). http://dx.doi.org/10.1016/j.watres.2015.08.003 0043-1354/© 2015 Elsevier Ltd. All rights reserved.

Heberer, 2002; Heberer et al., 2002; Lishman et al., 2006; Roberts and Thomas, 2006), little has been published on their ecotoxicological effects on aquatic organisms, especially marine invertebrates. Nevertheless, though the amount of pharmaceutical drugs released into the environment may be relatively low in comparison with other pollutants (e.g. pesticides and metals), their continuous input, their potential accumulation and/or their chronic ecotoxicity has already been demonstrated in non-target species (e.g. Besse and Garric, 2008; Escher and Fenner, 2011; Fent et al., 2006; Ferrari et al., 2003; Hernando et al., 2006; McEneff et al., 2014). The majority of these studies have been concentrated on acute studies but recent works have shown that the risk of acute toxicity is unlikely at measured environmental concentrations (Fent et al., 2006; Webb, 2004a, 2004b). Furthermore, due to their continuous presence at low concentrations in the aquatic environment, pharmaceuticals will most likely have chronic rather than acute toxic effects (Almeida et al., 2015; Crane et al., 2006; Fent et al., 2006; Freitas et al., 2015b). Therefore standardized acute tests may not be the most appropriate basis for the ecotoxicological hazard assessment of pharmaceuticals (Ferrari et al., 2004), and studies on chronic effects

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are lacking for most of them (Carlsson et al., 2006). The most common drugs and their metabolites found in the aquatic environment (mostly at the ng/L and mg/L level) include antiinflammatory drugs (ibuprofen, naproxen, diclofenac), lipid regulators (bezafibrate, gemfibrozil), anticonvulsion drugs (carbamazepine) and various antibiotics (trimethoprim, oxytetracycline, sulfamethoxazole) (Fent et al., 2006; Zhang et al., 2008). These compounds are mostly detected in the high ng/L to low mg/L range (Calisto et al., 2011a; Fatta-Kassinos et al., 2011; Fent et al., 2006; Heberer, 2002), concentrations that may be sufficient to induce toxic effects (Hernando et al., 2006). Carbamazepine (CBZ), an antiepileptic pharmaceutical compound and mood stabilizing drug, has attracted particular attention in recent years due to its widespread detection in aquatic environments (Heberer, 2002; Li, 2014; Santos et al., 2010; Ternes, 1998). Since CBZ has a low affinity for organic matter, sorption to sewage sludge is not an effective removal pathway, less than 10% of this drug is removed in standard WWTPs (Ternes, 1998; Zhang et al., 2008) and is therefore being detected in municipal wastewaters (Falås et al., 2012; Heberer, 2002; Ternes, 1998), surface waters (Andreozzi et al., 2003; Metcalfe et al., 2003; Ternes, 1998), and drinking waters (Fent et al., 2006; Heberer et al., 2004; Stackelberg et al., 2004, 2007; Togola and Budzinski, 2008). It is also considered to be highly resistant to biodegradation (Falås et al., 2012). Furthermore, CBZ is resistant to (photo-)chemical degradation in the environment, with a half-life in the order of 100 days in surface waters, and has therefore been suggested as a marker of wastewater input (Calisto et al., 2011b; Clara et al., 2004). Although there is currently no environmental regulation of CBZ internationally, its toxicity has been shown in algae, vertebrates and invertebrate organisms (Aguirre et al., 2006; Martínez et al., 2013a; Almeida et al., 2014, 2015; Gagne Li et al., 2010, 2009; Santos et al., 2010; Triebskorn et al., 2007; Tsiaka et al., 2013). Recent studies revealed acute and chronic effects of CBZ in marine invertebrates, with higher impacts resulting from chronic exposure tests (Ferrari et al., 2003; Oetken et al., 2005; Zhang et al., 2012). Furthermore, studies conducted with environmentally relevant concentrations have shown that CBZ can affect the biochemical and physiological performance of marine invertebrates (Almeida et al., 2015; Chen et al., 2014; Contardo-Jara et al., 2011; Martin-Diaz et al., 2009). Although an increasing number of studies have used bivalves as bioindicator species for pharmaceuticals contamination, including CBZ, very little is known on the impacts of these compounds on polychaetes. Usually polychaetes are the most abundant group in estuarine ecosystems (Rodrigues et al., 2011), being considered an important component of an ecotoxicological toolbox for sediment quality assessment, due to their abundance, ecological relevance and constant contact with contaminants in sediment and water column. Polychaetes have already shown to be highly useful bioindicators of toxicity induced by organic and inorganic contaminants (e.g. Durou  et al., 2009), organic matter et al., 2007; Freitas et al., 2012; Sole enrichment (Carregosa et al., 2014), or even climate change (Freitas et al., 2015c; Pires et al., 2015). Nevertheless, their responses to pharmaceutical drugs are largely unstudied (Freitas et al., 2015d; Maranho et al., 2015). Thus, the present study aimed to determine the effects of a chronic exposure (28 days) to environmentally relevant concentrations of CBZ (0.00e9.00 mg/L) in the clam Scrobicularia plana and in the polychaete Diopatra neapolitana. Toxicity of CBZ was evaluated through a battery of physiological and biochemical parameters. 2. Methodology 2.1. Study area and test organisms Scrobicularia plana (da Costa, 1778) and Diopatra neapolitana

(Delle Chiaje, 1841) specimens were collected in the Mira channel (Ria de Aveiro lagoon, Portugal) during Autumn to avoid the reproductive period of both species. In order to minimize the effect of body size on biochemical and physiological responses, organisms with similar size were used in the laboratory experiments: clams length varied between 52.7 and 47.3 mm, and between 36 and 42 mm width; polychaetes length, measured at 10th chaetiger, ranged between 6.95 and 7.35 mm. Sampling site was characterized by low organic matter content (<4%), redox potential of 19.4 mV, 1.5 mg/L of O2 and temperature of 16 ± 2
C. The Ria de Aveiro is composed by other channels, being the Mira channel considered as relatively pristine due to the low metal and metalloids concentrations (Castro et al., 2006; Freitas et al., 2014). Despite CBZ was detected by Calisto et al. (2011a) in surface water of the Ria de Aveiro, collected in a coastal touristic area as well as in wastewaters from two main WWTPs of Aveiro with concentrations ranging between 0.1 and 0.7 mg/L, this drug was not found at the Mira channel (Calisto et al., 2011a). For these reasons, the Mira channel was chosen as the present sampling area. In the present study the bivalve S. plana was selected since it has been considered a “key species” due of its ubiquity, abundance and  et al., 2009). It is a importance in the estuarine trophic chain (Sole natural resource of high importance not only as a food resource for avifauna and ichtyofauna, but also due to its economic value, especially in southern Portugal and Spain (Langston et al., 2007). This species has been largely used both in biomonitoring programs (Cheggour et al., 2005) and toxicity testing (among others Bergayou  et al., 2009; Boldina-Cosqueric et al., 2010; Coelho et al., 2014; Sole et al., 2009; Tankoua et al., 2013). D. neapolitana was selected for this study since it plays an important ecological role, by constituting an important food source for populations of birds, fish and other invertebrates (BIOREDE). At the Ria de Aveiro this species is economically important, as it is intensively collected by bait diggers to be used as fresh fish bait (Cunha et al., 2005). Recently, D. neapolitana showed to be sensitive to natural (Carregosa et al., 2014; Pires et al., 2015; Freitas et al., 2015c) and anthropogenic (Freitas et al., 2012, 2015d) changes in the aquatic environment. 2.2. Experiment set up In the field clams and polychaetes were collected from the same area to study the impact of CBZ. After sampling, clams and polychaetes (inside their tubes) were transported to the laboratory. In the laboratory D. neapolitana specimens were pushed out from their tubes, washed with natural seawater and placed in aquaria for acclimatization during 2 weeks. Both species were placed in different containers filled with a mixture of fine and medium sediment from the sampling area (approximately 1/3 of the height of the aquarium) and artificial seawater (salinity 25 g/L), with a photoperiod of 12 h light: 12 h dark, under constant temperature (18 ± 1
C), pH between 7.8 and 7.9, and to continuous aeration. Every other day, the clams were fed with AlgaMac Protein Plus, Aquafauna Bio-Marine, Inc (150 000 cells/animal) and the polychaetes were fed ad libitum with small fragments of frozen cockles (Pires et al., 2012). Salinity was set up by the addition of artificial sea salt (Tropic Marin® SEA SALT from Tropic Marine Center) to deionized water. After depuration, S. plana and D. neapolitana were exposed to environmental concentrations of CBZ (0.00; 0.30; 3.00; 6.00 and 9.00 mg/L) during 28 days. Exposure concentrations were prepared every week, by diluting a 10 mg/L stock solution of CBZ. This stock solution was maintained at 4
C and protected from sunlight. The selected range of concentrations was chosen according to concentrations currently found in aquatic systems worldwide (Li, 2014;

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Martin-Díaz et al., 2009). The first two concentrations (0.30 and 3.00 mg/L) are comparable with most of the measured environmental concentrations, while the two highest (6.00 and 9.00 mg/L) are comparable with the maximum levels revealed in aquatic ecosystems (for review see, Fent et al., 2006; Heberer, 2002). For each condition, 10 clams and 14 polychaetes were placed in the same aquarium filled with a mixture of fine and medium sediment from the sampling area (approximately 1/3 of the height of the aquarium) and artificial seawater (salinity 25 g/L). The organisms were submitted to constant temperature of 18 ± 1
C, pH between 7.8 and 7.9, a photoperiod of 12:12 h (light/dark) and continuous aeration. During experiments pH was continuously monitored and controlled by pHstat system (Aquamedic AT Controler). Salinity and temperature were monitored on a daily bases using specific probes (Hanna Instruments) for each parameter. During the experiment, water was renewed once a week and concentrations re-established. Animals were daily checked for mortality, and fed once every two days as referred previously. After 28 days of exposure clams and polychaetes were frozen for biochemical analyses. 2.3. Laboratory analyses From each condition, specimens that survived after exposure were pulverized with liquid nitrogen and divided into 0.5 g aliquots for carbamazepine quantification and biochemical analysis. Samples obtained were sonicated for 15 s at 4
C, centrifuged for 10 min at 10 000 g at 4
C and supernatant extracted with water, for CBZ quantification, or with a specific buffer for each analysis (1:2, w/v). 2.3.1. Carbamazepine quantification 2.3.1.1. Reagents. The polyclonal antibody against mouse (IgG F(c) domain, from goat, lot 20 185) and the anti-CBZ monoclonal antibody (mouse IgG1, clone B3212M, lot 5 K32007) were purchased from Acris Antibodies (Germany) and BIODESIGN International (Meridian Life Science Inc., USA), respectively. The tracer was produced and characterized as described in Bahlmann et al. (2009). 3,30 ,5,50 -Tetramethylbenzidine (puriss.), tetrabutylammonium borohydride (>97%, CAS number: 33725-74-5), sodium phosphate dibasic dihydrate (>99%), sodium phosphate monobasic dihydrate (>99%, CAS number: 10028-24-7), potassium sorbate (>99%, CAS number: 24634-61-5), potassium dihydrogen citrate (>99%, CAS number: 866-83-1), hydrogen peroxide (30%, CAS number: 772284-1) and Tween™20 (CAS number: 005-64-5) were purchased from Fluka. Ethylenediamine tetraacetic acid disodium salt dihydrate (>99%, CAS number: 60-00-4), and sodium chloride (99.5%, CAS number: 7647-14-5) were from Panreac. Dimethylacetamide (CAS number: 127-19-5), tris(hydroxymethyl) aminomethane (TRIS, p.a.) and glycine (99.8%, CAS number: 56-40-6) were purchased from VWR Prolabo. 2.3.1.2. Immunoassay procedure. Carbamazepine was quantified in supernatants obtained using a water extraction (1:2 w/v) (Almeida et al., 2014). Since during the experiment the water in the containers was renewed and CBZ concentrations were re-established every week, to assess if the contamination levels were maintained constant in the assay, aliquots of water were collected for quantification at day 0 (beginning of the assay) and at day 28 (which corresponds to 7 days after the last medium renewal), from the water where the organisms were exposed and from blanks (sediment-seawater containers spiked with CBZ with the same concentrations used for the exposure but with no organisms). To assess the behaviour of CBZ in the exposure medium, prior to the experiments, blanks with sediment-seawater and blanks with

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only seawater, both with no organisms, were exposed to CBZ (3.00 mg/L) during 5 days. Results obtained showed that more that 91% of the CBZ was maintained in the water column in both blanks (seawater and sediment-seawater), not differing significantly from each other. Thus, it was concluded that CBZ did not adsorbed significantly in the sediment during the exposure, allowing conducting the assay in sediment-seawater and without the necessity to quantify the CBZ in the sediment. To evaluate possible matrix effects on CBZ quantification in clams and polychaetes, a quality control was performed. For this, supernatants (an aqueous extract) were obtained from noncontaminated specimens by extraction with deionized water (1:2 w/v). Samples were spiked with a known concentration of carbamazepine: 0.30, 3.00 mg/L and quantified to assess the recovery percentage. Direct competitive ELISA assay (Enzyme-Linked Immunosorbent Assay), developed by Bahlmann et al. (2009) and modified by Almeida et al. (2014) was applied to quantify CBZ. This assay is based on a direct competition of the analyte (CBZ) and a so-called tracer (CBZ analogue linked to the enzyme horseradish peroxidase) for the primary antibody's (monoclonal antibody against CBZ) binding sites. 96-well high-binding microtiter plates were coated with a polyclonal antibody against mouse IgG (1 mg/L, 200 mL per well) diluted in phosphate buffered saline (PBS) (10 mM sodium dihydrogen phosphate, 70 mM sodium hydrogen phosphate, 145 mM sodium chloride, pH 7.6). Plates were covered with Parafilm® to prevent evaporation and incubated overnight (approximately 16e18 h) on a Titramax 100 plate shaker at 900 rpm. After overnight incubation, plates were washed three times with PBS containing 0.05% (v/v) TweenTM 20 (PBS-T) using an automatic 8channel plate washer. Monoclonal antibody against CBZ was also diluted in PBS (7.61  10-5 mg/mL, 200 mL per well), added to wells and incubated for 1 h. Then a three-cycle washing step of the plate was performed again. After that, 50 mL of tracer solution (147 pmol/ L in sample buffer) and 150 mL of CBZ standard solutions or samples were added per well and incubated for 30 min. Sample buffer consisted of 1 M glycine, 3 M sodium chloride and 2% (w/v) of EDTA, pH 9.5. After another three-cycle washing step (PBS-T), 200 mL of substrate solution was added per well and incubated for 30 min. Substrate solution consisted of 540 mL TMB-based solution (41 mM 3,30 ,5,50 -tetramethylbenzidine (TMB)), 8 mM tetrabutylammonium borohydride (TBABH) prepared in dimethylacetamide (DMA) under nitrogen atmosphere) in 21.5 mL of substrate buffer (220 mM citric acid, 0.66 mM sorbic acid potassium salt and 3 mM hydrogen peroxide). The TMB solution was freshly prepared for each run. The enzyme reaction was stopped by the addition of 1 M sulphuric acid (100 mL per well) and the optical density was read on a microplate spectrophotometer at 450 nm and referenced to 650 nm. Data was analysed using SoftMax® Pro Software (version 5.3, Molecular Devices). All samples and standards were determined in triplicate on each plate. A four-parametric logistic equation (4 PL) was fitted to the standards' mean values in order to obtain the calibration function (Findlay and Dillard, 2007). A calibration curve was established for each plate using eight calibrators, with concentrations between 0 and 100 mg/L for a better convergence of the curve fitting. For the analysis of clam's supernatants standards were prepared, in ultrapure water, by diluting a 10 mg/L stock solution of CBZ (also prepared in ultrapure water). For the analysis of water samples (blanks) the standards were prepared in seawater (25 g/L NaCl) by diluting a stock solution of CBZ of the same concentration. 2.3.2. Biochemical parameters 2.3.2.1. Reagents. Trichloroacetic acid (>98%, CAS number: 76-039), sodium dihydrogen phosphate monohydrate (99.0e102.0%, CAS

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number: 10049-21-5), disodium hydrogen phosphate dehydrate (99.0%, CAS number: 10028-24-7) and Triton X-100 (CAS number: 9036-19-5) was purchased from Merck Millipore. Polyvinylpyrrolidone (CAS number: 9003-39-8), bovine serum albumin (for molecular biology, CAS number: 9048-46-8), reduced glutathione (98.0%, CAS number: 70-18-8), superoxide dismutase (CAS number: 9054-89-1), formaldehyde solution (for molecular biology, CAS number: 50-00-0) and sulfosalicylic acid (>99%, CAS number: 5965-83-3) were purchase from SigmaeAldrich. Dithiothreitol (CAS number: 3483-12-3) was purchase to VWR Prolabo. Dipotassium phosphate (>99%, CAS number: 7758-11-4), potassium dihydrogen phosphate (98.0e100.5%, CAS number: 7778-770) and oxidized glutathione (CAS number: 27025-41-8) were purchased from Panreac Applichem.

(2014). Standards of formaldehyde (0e150 mM) were prepared. Absorbance was read at 540 nm. CAT activity was expressed in U per g of FW. One unit (U) is defined as the amount of enzyme that caused the formation of 1.0 nmol of formaldehyde per min, under assay conditions.

2.3.2.2. Biochemical procedures. For each biochemical parameter 0.5 g of tissue per individual was used. Extraction was performed with specific buffers to determine: glycogen content (GLYC), protein content (PROT), lipid peroxidation (LPO), reduced (GSH) and oxidized (GSSG) glutathione content, superoxide dismutase (SOD), catalase (CAT) and glutathione S-transferases (GSTs) activities. All biochemical parameters were performed in duplicate, using five organisms each. For LPO supernatants were extracted using 20% (v/ v) trichloroacetic acid (TCA). For CAT, SOD, GST, GLYC and PROT content the extraction was done with sodium phosphate buffer (50 mM sodium dihydrogen phosphate monohydrate; 50 mM disodium hydrogen phosphate dihydrate; 1 mM ethylenediamine tetraacetic acid disodium salt dihydrate (EDTA); 1% (v/v) Triton X100; 1% (v/v) polyvinylpyrrolidone (PVP); 1 mM dithiothreitol (DTT)). GSH and GSSG were determined using 0.6% sulfosalicylic acid in potassium phosphate buffer (0.1 M dipotassium phosphate; 0.1 M potassium dihydrogen phosphate; 5 mM EDTA; 0.1% (v/v) Triton X-100; pH 7.5).

2.4. Data analysis

2.3.2.3. Glycogen and proteins content. Glycogen (GLYC) was quantified according to sulphuric acid method (Dubois et al., 1956), using glucose standards (0e2 mg/mL). Absorbance was measured at 492 nm. Concentration of glycogen was expressed in mg per g of fresh weight (FW). Protein content (PROT) was determined following the spectrophotometric method of Biuret (Robinson and Hogden, 1940), using bovine serum albumin (BSA) as standards (0e40 mg/mL). Absorbance was read at 540 nm. Results were expressed in mg per g of FW. 2.3.2.4. Indicators of cellular damage. Lipid Peroxidation (LPO) was measured according to Ohkawa et al. (1979) and the modifications referred by Carregosa et al. (2014). Absorbance was read at 532 nm (ε ¼ 156 mM1 cm1). LPO was expressed in nmol of MDA formed per g of FW. Reduced (GSH) and oxidized (GSSG) glutathione content were determined according to Rahman et al. (2007), using reduced and oxidized glutathione standards (0e60 mmol/L). Absorbance was measured at 412 nm. Results were expressed as nmol per g of FW. GSH/GSSG ratio was determined. 2.3.2.5. Antioxidant enzymes. Superoxide Dismutase activity (SOD) was determined based on the method of Beauchamp and Fridovich (1971), with some adaptations (Carregosa et al., 2014). Standards of SOD were prepared (0.25e60 U/mL). SOD activity was measured spectrophotometrically at 560 nm. SOD was expressed in U per g of FW where U represents the quantity of enzyme that catalyses the conversion of 1 mmol of substrate per min. Catalase activity (CAT) was quantified according Johansson and Borg (1988) and to modifications performed by Carregosa et al.

2.3.2.6. Biotransformation enzymes. The activity of glutathione Stransferases (GSTs) was determined according to Habig et al. (1974) and the adaptations described by Carregosa et al. (2014). GSTs activity was measured spectrophotometrically at 340 nm (ε ¼ 9.6 mM1 cm1). The enzymatic activity was expressed in U/g of FW. One unit of enzyme is defined as the amount of enzyme that causes the formation of 1 mmol of dinitrophenyl thioether per min under assay conditions.

Bioconcentration Factor (BCF) was determined dividing the concentration of CBZ present in organism's tissues by the spiked CBZ concentration for each exposure condition (Gobas and Morrison, 2000). Biochemical descriptors and CBZ quantification data were submitted to hypothesis testing using permutation multivariate analysis of variance with the PERMANOVA þ add-on in PRIMER v6 (PRIMER-E Ltd, UK, Anderson et al., 2008). A one-way hierarchical design, with CBZ exposure concentration as the main fixed factor, was followed in this analysis. Pseudo-F values in the PERMANOVA main tests were evaluated in terms of significance (Anderson et al., 2008). When the main test revealed statistical significant differences (p  0.05) pairwise comparisons were performed. In the pairwise comparisons t-statistic was evaluated in terms of significance. Values lower than 0.05 were considered as significantly different. Taking into account the mortality rate, 6 individuals were used per descriptor. The null hypotheses tested were: a) for each exposure concentration, no significant differences exist between species; b) for each species, no significant differences exist among exposure concentrations. For each species, significant differences among exposure concentration are represented with letters. Significant differences between species at each concentration are represented with an asterisk. Matrix with biomarkers responses per exposure concentration was used to calculate the Euclidean distance similarity matrix. This matrix was simplified through the calculation of the distance among centroids matrix based on the 5 exposure conditions, which was then submitted to ordination analysis, performed by Principal Coordinates (PCO). Pearson correlation vectors (r > 0.75) of biomarkers responses descriptors were provided as supplementary variables being superimposed on the PCO graph. 3. Results 3.1. Mortality At the end of the chronic assay mortality was observed at 0.30 mg/L for S. plana (10%). For D. neapolitana mortality was recorded at control and 0.30 mg/L (12.5% each), and at 3.00, 6.00 and 9.00 mg/L (25% each). 3.2. Carbamazepine quantification CBZ concentration measured in water and blank samples are presented in Table 1. Results reveal that, for each condition, the CBZ concentration did not vary significantly along the exposure period. Similar results were found in samples from blanks, with no CBZ losses along the exposure period, which indicated that no

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photodegradation or adsorption onto the aquaria occurred. Results obtained for the quality control (Table 2) indicate that in the aqueous extracts (supernatants) obtained from noncontaminated clams and polychaetes by extraction with deionized water and spiked with CBZ (0.30 and 3.00 mg/L), all the concentrations spiked could be quantified by ELISA, revealing that no matrix interferences were occurring. Results for CBZ quantification in organism's tissues (Fig. 1A) revealed that both species increased the CBZ concentration along exposure gradient. Comparing both species, polychaetes accumulated higher amounts of CBZ in each of the tested conditions than clams, with significant differences at the lowest and the highest tested concentrations (cf. Fig. 1A). Bioconcentration factor (BCF) determined for each species at each condition is presented in Fig. 1B. Results obtained showed that both species tended to bioconcentrate more CBZ at the lowest exposure concentration. With increasing exposure gradient both species tended to limit the uptake of CBZ, especially S. plana, demonstrated by the lower BCF values found at concentrations 3.00, 6.00 and 9.00 mg/L. 3.3. Biochemical parameters 3.3.1. Protein and glycogen content Protein (PROT) content in D. neapolitana was significantly lower in organisms exposed to 0.30, 3.00 and 6.00 mg/L compared to the control and the highest CBZ concentration (9.00 mg/L) (Fig. 2A). In clams a similar trend was observed but significant differences were only noticed between control and concentrations of 0.30 and 6.00 mg/L (cf. Fig. 2A). At all conditions, D. neapolitana presented significantly higher protein content than S. plana. (cf. Fig. 2A). As for protein content, glycogen (GLYC) in D. neapolitana tended to be lower at 0.30, 3.00 and 6.00 mg/L, with significant differences between control and 3.00 mg/L of CBZ (Fig. 2B). The highest glycogen content was found at control and at the highest exposure concentration (9.00 mg/L) (cf. Fig. 2B). For S. plana a similar pattern was found, with significantly lower values at the lowest concentrations (0.30, 3.00 and 6.00 mg/L) and higher values at the highest concentration where no significant differences were found relative to control (cf. Fig. 2B). At each exposure concentration significant differences were found between species, with D. neapolitana presenting the highest values in all the tested conditions (cf. Fig. 2B). 3.3.2. Indicators of cellular damage Results obtained showed that D. neapolitana presented significantly higher LPO at 0.30 and 9.00 mg/L of CBZ (Fig. 3A). For S. plana, LPO significantly increased along the increasing exposure gradient, with the highest values at concentration 9.00 mg/L (cf. Fig. 3A). Comparing both species, significant differences were found in control, 0.30 and 3.00 mg/L conditions (cf. Fig. 3A). In control and at the lowest CBZ concentration D. neapolitana presented higher LPO, while for the remaining conditions S. plana presented the highest values (cf. Fig. 3A).

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The ratio between reduced (GSH) and oxidized (GSSG) glutathione showed that both species presented significantly lower GSH/ GSSG values when under CBZ compared to the control (Fig. 3B). D. neapolitana significantly decreased GSH/GSSG ratio along the exposure gradient, with the lowest values at the highest concentration. Clams presented a similar pattern, but less pronounced. When comparing both species it is possible to observe that clams presented significantly higher GSH/GSSG values in all the conditions tested. 3.3.3. Antioxidant enzymes Results obtained for SOD revealed that both species tend to increase the activity of this enzyme with the increasing CBZ concentration (Fig. 4A). However, at the highest CBZ concentration (9.00 mg/L) for D. neapolitana and at the two highest concentrations (6.00 and 9.00 mg/L) for S. plana the activity of this enzyme significantly decreased, showing no significant differences to control (cf. Fig. 4A). When comparing both species it is possible to observe that at the two highest CBZ concentrations D. neapolitana presented higher SOD values than S. plana (with significant differences at 6.00 mg/L), but at the two lowest concentrations (0.30 and 3.00 mg/ L), although not significantly different, S. plana presented higher SOD activity than D. neapolitana (cf. Fig. 4A). The activity of CAT was maintained along the exposure gradient in D. neapolitana, with no significant differences among conditions (Fig. 4B). In S. plana a significant increase, compared to control, was observed in the activity of CAT at concentrations 0.30 and 3.00 mg/L (cf. Fig. 4B). Comparing D. neapolitana and S. plana, the polychaete presented significantly higher CAT activity in all the tested conditions (cf. Fig. 4B). 3.3.4. Biotransformation enzymes Concerning the activity of GSTs, along the exposure gradient, D. neapolitana presented significantly lower values at concentrations 3.00 and 6.00 mg/L, while S. plana revealed no significant differences among the tested conditions (Fig. 4C). S. plana presented higher GSTs activity than D. neapolitana, with significant differences at all tested concentrations. 3.4. Multivariate analysis In Fig. 5 centroid PCO ordination graph is represented resulting from applying a multivariate analysis to biochemical parameters. The PCO1 axis explained 59.9% of the total variation of data, clearly separating both species (S. plana at the negative side of the axis and D. neapolitana at the positive side of the axis). PCO2 axis explained 19.4% of the total variation. In S. plana this axis separates the highest concentrations (6.00 and 9.00 mg/L, in the positive side) from the remaining ones (CTL, 0.30, 3.00 mg/L, in the negative side), while for N. neapolitana concentrations 3.00 and 6.00 mg/L are in the negative pole and the remaining conditions (CTL, 0.30 and 9.00 mg/L) appear in axis positive pole. From PCO analysis it is possible to observe that: i) S. plana control specimens are strongly

Table 1 CBZ concentration (mg/L), determined by ELISA, in water where the organisms were exposed to the chronic assay, in blanks at the beginning (day 0) and at the end of the assay (day 28). Significant differences (p  0.05) between exposure periods for CBZ in water from the chronic assay and in blanks are presented with letters (a).
Water (mg/L) e day 0

Water (mg/L) e day 28

Blanks (mg/L)- day 0

Blanks (mg/L) e day 28

0.00 0.30 3.00 6.00 9.00





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Table 2 CBZ concentration (mg/L), determined by ELISA, for the quality control test and the respective recovery percentage. Significant differences (p  0.05) between exposure concentrations (spiked and determined by ELISA) are presented with letters (a). Spiked concentration (mg/L)

ELISA response (mg/L) D. neapolitana

0.30 3.00

0.34 (0.01) 3.4 (0.3)a

a

correlated with GSH/GSSG descriptor, since in this condition this species presented the highest GSH/GSSG value; ii) organisms of S. plana exposed to 6.00 and 9.00 mg/L showed a strong correlation with GSTs, because organisms exposed to these concentrations presented the highest activity of this group of enzymes; iii) glycogen and protein content are strongly correlated with D. neapolitana individuals under control and 9.00 mg/L, since at these conditions glycogen and protein presented the highest content; iv) individuals of D. neapolitana exposed to 3.00 and 6.00 mg/L were well correlated with SOD because at these concentrations the activity of this enzyme was significantly enhanced. 4. Discussion Species Scrobicularia plana and Diopatra neapolitana have been widely used to assess effects of contaminants in the aquatic

Fig. 1. A- Carbamazepine (CBZ) concentration and B- CBZ Bioconcentration factor (BCF) in S. plana and D. neapolitana when exposed to increasing concentrations of CBZ during 28 days. Values are the mean ± STDEV of six replicates. For each species, significant differences (p  0.05) among exposure concentrations are presented with letters (aed). At each condition, significant differences between species are represented by an asterisk.

Recoveries (%) S. plana 0.33 (0.02) 2.4 (0.2)a

a

D. neapolitana

S. plana

112 (4) 113 (11)

111 (7) 81 (6)

ecosystems (among others, Ahmad et al., 2011; Boldina-Cosqueric  et al., 2009; Tankoua et al., et al., 2010; Silva et al., 2012; Sole 2013 Freitas et al., 2012; Freitas et al., 2015a, b, d) but little is known regarding their tolerance to pharmaceutical drugs, namely carbamazepine. Thus, the present work evaluated the effects of CBZ, exposing polychaetes and bivalves to environmentally relevant concentrations for 28 days (chronic assay) as most of literature on pharmaceutical drug impacts on marine invertebrates is based on acute exposures (Quinn et al., 2008; Yang et al., 2008) that are not sufficient to allow for a potential acclimatization to a new environment or may not induce responses. Since the majority of studies reporting toxic effects caused by pharmaceuticals use concentrations that are not likely to occur in the environment, CBZ concentrations tested were chosen as representative of those detected in

Fig. 2. A- Protein (PROT) and B-glycogen (GLY) content in S. plana and D. neapolitana when exposed to increasing concentrations of CBZ during 28 days. Values are the mean ± STDEV of six replicates. For each species, significant differences (p  0.05) among exposure concentrations are presented with letters (aeb). At each condition, significant differences between species are represented by an asterisk.

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Fig. 3. Indicators of oxidative stress: A-lipid peroxidation (LPO) and B-ratio between reduced and oxidized glutathione (GSH/GSSG) in S. plana and D. neapolitana when exposed to increasing concentrations of CBZ during 28 days. Values are the mean ± STDEV of six replicates. For each species, significant differences (p  0.05) among exposure concentrations are presented with letters (aec). At each condition, significant differences between species are represented by an asterisk.

the aquatic environment. Furthermore, this study also intended to understand how different species, under the same conditions, respond to the same stressors. For this, oxidative stress related biomarkers were evaluated in bivalves and clams. The ability of organisms to accumulate CBZ (under laboratory or field conditions) has been reported in recent studies (Contardo-Jara et al., 2011; Garcia et al., 2012; Martínez-Bueno et al., 2013; Ramirez et al., 2007; Vernouillet et al., 2010). In the present study, CBZ uptake increased with the increase of exposure concentration in both species, demonstrating the risk of CBZ bioaccumulation in polychaetes and clams. Regarding Bioconcentration Factor (BCF), determined for both species, the highest values were found at CBZ 0.30 mg/L, with significantly lower values at the remaining exposure concentrations. These results may indicate the effort of both species to prevent CBZ accumulation. Regarding clams, these results may confirm the capacity of S. plana to decrease filtration rates in increasingly stressful conditions; i.e. higher CBZ concentrations, by closing their valves. Gosling (2003) reported that bivalves can isolate their tissues from the external environment by closing their valves thus, protecting themselves against contaminants. In studies performed by Contardo-Jara et al. (2011) the highest BCF values were also found at the lowest CBZ concentrations. The closure of valves in the presence of CBZ was also observed by Chen et al. (2014) when submitting the freshwater clam Corbicula fluminea to 5 and 50 mg/L of CBZ, during 30 days, indicating that exposure to environmentally

Fig. 4. Antioxidant and biotransformation enzymes: A- superoxide dismutase (SOD); B- catalase (CAT) and C-glutathione S-transferases (GSTs) in S. plana and D. neapolitana when exposed to increasing concentrations of CBZ during 28 days. Values are the mean ± STDEV of six replicates. For each species, significant differences (p  0.05) among exposure concentrations are presented with letters (aec). At each condition, significant differences between species are represented by an asterisk.

relevant concentrations is enough to alter siphoning behaviour in these clams. Nevertheless, an opposite trend was demonstrated by  et al. (2010) evaluating the effects of propranolol and acetSole aminophen on the feeding rate of mussels (Mytilus galloprovincialis). Concerning D. neapolitana, the lower accumulation rate of CBZ at the highest exposure concentration may result from the lower feeding rate observed in polychaetes when under stressful conditions. Moreira et al. (2005, 2006) demonstrated that Hediste diversicolor significantly decreased feeding capacity after exposure to contaminated sediments and that the more contaminated the sediments were, the lower was the feeding rate. This behaviour

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Fig. 5. Biochemical responses of S. plana and D. neapolitana exposed to increasing CBZ concentrations during 28 days, plotted on axes 1 and 2 of a Principal Coordinates (PCO) graph. Biochemical responses are superimposed on the PCO (r > 0.75). The control (CTL) and concentrations of exposure (0.03, 0.30, 3.00, 9.00 mg/L) are indicated. The biomarkers presented are: GLYC, glycogen content; PROT, protein content; GSH/GSSG, ratio between reduced and oxidized glutathione; GSTs, glutathione S-tansferases; LPO, lipid peroxidation; CAT, catalase; and SOD, superoxide dismutase.

may explain why D. neapolitana presented a lower BCF at higher CBZ concentrations. When assessing the effects of CBZ, our findings revealed that both polychaete and the bivalve decreased their glycogen and protein content at concentrations 0.30, 3.00 and 6.00 mg/L. At these concentrations D. neapolitana decreased between 43 and 35% protein and between 20 and 13% glycogen content, while for S. plana protein content decreased between 25 and 21% and glycogen content decreased between 30 and 36%. Nevertheless, no significant differences were found between control and the highest CBZ concentration (9.00 mg/L) in both species, which may be related to the lowest bioaccumulation rate observed at this condition since both species may be able to activate their defence response by reducing their feeding capacity. Almeida et al. (2015) also found a significant glycogen increase along the CBZ concentration gradient, especially at the highest concentration (9.00 mg/L), in Ruditapes philippinarum exposed to CBZ for 28 days, which was correlated to the lowest clearance rate observed in clams exposed to this concentration. Duquesne et al. (2004) observed that the reduction in clearance rate of the bivalve Macoma balthica, as an attempt to limit exposure to contamination, led to a slowdown in metabolism resulting in a relatively stable glycogen concentration. BoldinaCosqueric et al. (2010), evaluating the response of S. plana to a pollution gradient, revealed that no significant differences were found on glycogen content between clams from distinct locations. Studies conducted with R. philippinarum, R. decussata and R. corrugata showed that these species maintained or increased glycogen content when exposed to stressful salinities (Carregosa et al., 2014a). Concerning polychaetes, recent studies conducted by Maranho et al. (2015) also showed that energy reserves, assessed by storage of total lipids, were positively correlated with the increasing CBZ concentrations in the polychaete Hediste diversicolor. Studies conducted by Carregosa et al. (2014b) revealed that D. neapolitana increased protein and glycogen content when under stressful organic matter enrichment conditions. Cellular damages induced by CBZ on polychaetes and clams were assessed by measuring LPO in both species. Our data revealed cellular damages induced by CBZ, showing that this drug induced LPO in both species, with stronger impacts in S. plana which

augmented LPO between 8 (at 0.30 mg/L) and more than 100% (at 9.00 mg/L), while D. neapolitana showed an LPO increase between 5 (at 6.00 mg/L) and 27% (at 9.00 mg/L). Damages provoked by CBZ were also noticed by the decrease in the GSH/GSSG ratio since both species strongly decreased the GSH/GSSG values along the increasing CBZ concentration gradient, with the lowest values at the highest concentration. Previous studies demonstrated that pharmaceutical drugs (including CBZ) induced oxidative stress in marine invertebrates, including bivalves and polychaetes. Recently, Freitas et al. (2015a, b) showed that acute and chronic exposure to CBZ (3.00 mg/L) increased the LPO in S. plana. Chen et al. (2014) also reported that MDA significantly increased in gills and digestive gland of C. fluminea with the increased of CBZ concentrations. Martin-Diaz et al. (2009) and Tsiaka et al. (2013) also showed a significant increase in LPO in M. galloprovincialis when mussels were exposed to CBZ concentrations. Also Schmidt et al. (2011) showed that the pharmaceutical drug Diclofenac significantly induced LPO in Mytillus spp. after 96 h exposure. Regarding polychaetes, Maranho et al. (2014) demonstrated that LPO significantly increased with concentration of CBZ in the H. diversicolor, after a 14 days exposure period. Carregosa et al. (2014) demonstrated that LPO increased in D. neapolitana along increasing organic matter content in sediments. Moreira et al. (2006) showed that H. diversicolor presented higher lipid peroxide and lower GSH/GSSG values at the most contaminated sediments. In our study we measured the activity of CAT and SOD enzymes that are responsible for neutralizing ROS (reactive oxygen species) before they initiate radical chain reactions leading to lipid peroxidation. Results obtained revealed that, in both species, CAT and SOD were not able to act efficiently as antioxidant defender, which was at least partially responsible for increased LPO in clams and polychaetes exposed to CBZ. Except for D. neapolitana at the lowest CBZ exposure concentration, the activity of SOD increased with the increase of CBZ. However, at 9.00 mg/L for D. neapolitana and 6.00 and 9.00 mg/L for S. plana the activity of SOD decreased compared to the values obtained at lower concentrations, although presenting higher SOD activity values than control. This fact may indicate that SOD also seemed to be actively involved in ROS elimination at lower concentrations, but at higher CBZ concentrations this enzyme is not

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able to prevent oxidative stress which leads to an induction of LPO in both species at higher CBZ concentrations. The capacity of SOD to prevent LPO at lower CBZ concentrations is also demonstrated by the fact that at the lowest CBZ concentration D. neapolitana increased LPO which may result from the lack of increase of this enzyme's activity. Concerning CAT, in general, both species tended to maintain the activity of this antioxidant enzyme along the CBZ gradient. These results may indicate that in both species the hydrogen peroxide produced by SOD is not being consumed by CAT, especially at the highest CBZ concentrations. A similar pattern was found by Almeida et al. (2015) for R. philippinarum exposed to CBZ for 28 days. These authors showed that SOD activity increased when clams were exposed to CBZ, but at the highest CBZ exposure concentration the induction on the SOD activity was the lowest and CAT activity was similar to control. Freitas et al. (2015a, b) demonstrated that S. plana maintained or even decreased CAT and SOD activities when clams were exposed to CBZ (3.00 mg/L) during 96 h and 28 days. Also Contardo-Jara et al. (2011) found a decrease in the activity of SOD at the highest CBZ concentration (236 mg/L) in Dreissena polymorpha, after 7 days of exposure. Studies conducted by Li et al. (2009) showed that SOD activity decreased along the concentration gradient of CBZ (1 mg/L, 0.2 mg/L and 2 mg/L) after a 21 days exposure of rainbow trout Oncorhynchus mykiss. After a prolonged exposure (42 days), a strong inhibition of SOD activity was observed and attributed to the overproduction of ROS and the relatively low activity of antioxidant system. Moreover Milan et al. (2013) demonstrated that the SOD activity decreased markedly in R. philippinarum exposed to ibuprofen, suggesting that the contaminant reduced the antioxidant defences of this clam. Studies conducted by Gonzalez-Rey and Bebianno (2012) demonstrated an inhibition tendency in CAT activity over time in the mussel Mytilus galloprovinciallis exposed to the non-steroidal antiinflammatory drug ibuprofen for 15 days. Martin-Diaz et al. (2009) reported no significant alteration on CAT activity measured in gills of M. galloprovincialis exposed to CBZ for 7 days. On the other hand, Chen et al. (2014) observed that the increase in CAT activity after exposure of the clam C. fluminea to CBZ (5 and 50 mg/L, 30 days). Also Parolini et al. (2010) demonstrated that CAT activity reached values about two-fold higher than the control at the end of exposure (96 h) at the higher concentration of paracetamol. Although, to our knowledge no studies evaluated effects provoked in CAT and SOD enzymes in polychaetes after exposure to pharmaceutical drugs, Maranho et al. (2014) demonstrated that no significant changes were observed in the activity of the antioxidant enzymes glutathione peroxidase (GPX) and glutathione reductase (GR) in H. diversicolor exposed to an increasing concentration gradient of CBZ. Nevertheless, studies with D. neapolitana showed that this species was able to induce SOD and CAT activity when organisms were under organic matter enrichment (Carregosa et al., 2014c). Moreira et al. (2006) also demonstrated that H. diversicolor increased SOD and CAT activity when individuals were exposed to metal contaminated sediments. GSTs, an important enzymatic family of phase II of the biotransformation process, catalyse the conjunction of GSH with a wide group of compounds bearing electrophilic centres, playing an important role in the detoxification and excretion of endogenous compounds in invertebrates, including organic compounds, such as PAHs and PCBs and products of oxidative stress (Clark, 1989). GSTs isoenzymes are capable of inactivating lipoperoxidation products, such as lipid hydroperoxides (Sturve et al., 2008) by the use of GSH as a reducing agent that is oxidized to GSSG when peroxides are reduced (Contardo-Jara et al., 2010). These enzymes are also involved in the formation of thiol metabolites resulting from CBZ oxidation, in which GSH is conjugated with the carbonyl group of CBZ (Vernouillet et al., 2010). In the present study the GSTs activity

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was differently affected in polychaetes and clams. Results presented here showed that in S. plana GSTs activity did not change significantly along the CBZ gradient, while in D. neaplitana, except for the lowest concentration (0.30 and 9.00 mg/L), GSTs were inhibited (25% at 6.00 mg/L to 42% at 3.00 mg/L). The pattern observed in D. neapolitana and S. plana may explain the increase in lipid peroxidation recorded in both species. Also Almeida et al. (2015) demonstrated that the activity of this group of enzymes was not altered in R. philippinarum exposed to CBZ for 28 days. Freitas et al. (2015a) showed that S. plana from a non-polluted area did not significantly change GSTs activity when exposed to CBZ (3.00 mg/L) but increased their activity after a 28 days exposure (Freitas et al., 2015b). Several authors also showed an induction in GSTs activity in organisms exposed to different pharmaceutical drugs. Aguirre-Martínez et al. (2013b) and Martin-Diaz et al. (2009), revealed an induction in GSTs activity in the crab Carcinus maenas and in the mussel M. galloprovincalis exposed to CBZ. Antunes et al. (2013) also showed a significant increase of GSTs when R. philippinarum was exposed to paracetamol. Maranho et al. (2014) showed that GSTs were induced at higher CBZ concentrations in H. diversicolor. Also Carregosa et al. (2014c) demonstrated that D. neapolitana increased GSTs activity with the increase of organic matter content. Although similar among the exposure concentrations, our findings suggest that the higher GSTs activity recorded in S. plana may, in part, explain the lower CBZ accumulation by this species compared with D. neapolitana. 5. Conclusions Most of the studies aiming to evaluate effects of pharmaceutical drugs in marine invertebrates have been conducted with bivalves and very few with polychaetes. Since polychaetes are considered as a suitable bioindicator for the toxicity assessment of different types of contaminants (e.g. metals, organic enrichment), our findings can be applied to better predict long-term impacts of pharmaceuticals in marine and estuarine environments. Furthermore, because polychaetes are often the most abundant group of macroinvertebrates in marine and estuarine environments, we can assume that benthic communities exposed to environmental concentrations of pharmaceuticals could be in ecological risk. Overall the present study demonstrated that the chronic exposure of D. neapolitana and S. plana to CBZ will significantly influence their physiological and biochemical behaviour. Although our data may indicate that both species showed the ability to reduce accumulation along a CBZ exposure gradient, probably by reducing their feeding rates, our findings revealed that CBZ induced oxidative stress in D. neapolitana and S. plana. The cellular damage, assessed by LPO and GSH/GSSG, was induced in both species which were not able to efficiently develop antioxidant defences to prevent oxidative injuries. When comparing the effects provoked in both species it seems that S. plana is the most sensitive species, especially noticed by higher LPO values recorded in S. plana than in D. neapolitana; and lower antioxidant capacity especially noticed by lower SOD activity at the highest CBZ concentrations (6.00 and 9.00 mg/L). Acknowledgements This work was supported by the Portuguese Science Foundation (FCT) through CESAM: UID/AMB/50017/2013. Rosa Freitas and V^ ania Calisto benefited from post-doc grants (SFRH/BPD/92258/ 2013 and SFRH/BPD/78645/2011, respectively) given by the FCT. tia Velez benefited from a PhD grant (SFRH/BD/86356/2012) Ca given by the FCT.

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