Bioaccumulation Patterns, Element Partitioning and

Bioaccumulation Patterns, Element Partitioning and Biochemical Performance of Venerupis corrugata from a Low Contaminated System Catia Velez, Rosa Freitas, Amadeu Soares, Etelvina Figueira Departmento de Biologia & CESAM, Universidade de Aveiro, 3810-193 Aveiro, Portugal

Received 20 August 2014; revised 28 October 2014; accepted 31 October 2014 ABSTRACT: The current study reports metals and arsenic (As) concentrations present in sediments and in the native clam Venerupis corrugata, collected in the Ria de Aveiro, one of the most important aquatic systems of the Portuguese coast with high biodiversity and socio-economic value. Because of its ecological importance in its habitat, and being one of the most exploited bivalve mollusks in Portugal, several biochemical biomarkers were evaluated in order to illustrate the species status when under environmental conditions. The concentration of metals and As in the sediments showed an increase of contamination among areas (areas A–E). The results proved higher bioaccumulation in organisms from the area less contaminated (area A, BAF > 1). The concentration of metals and As was predominant (63.4%) in the insoluble fraction of clams. The biochemical evaluation evidenced an increase of oxidative stress in organisms from the most (D and E) and the less (area A) contaminated areas, demonstrated by higher LPO levels, CAT, and GSHt activities at these areas and the increase of methalotionines (MTs) along the concentration gradient. This suggests a preventive mechanism in order to protect cells against pollutants (metals and C 2014 Wiley Periodicals, Inc. Environ Toxicol 00: 000–000, 2014. As). V Keywords: clams; biomarkers; element partitioning; oxidative stress

INTRODUCTION Estuarine systems are exposed to a large amount of contaminants from natural and anthropogenic sources, such as erosion, volcanic eruptions, industry, agriculture, urbanization, and sewage discharges (Bergayou et al., 2009; Freitas et al., 2012a,b; Moschino et al., 2012; Spada et al., 2012). Nevertheless, estuaries are among the most relevant and dynamic ecosystems in terms of ecology and biodiversity (Rodrigues et al., 2009; Santos et al., 2014). Correspondence to: R. Freitas; e-mail: [email protected] Contract grant sponsor: National Funds through the Portuguese Science Foundation (FCT) Contract grant number: SFRH/BD/86356/2012 Contract grant sponsor: European Funds through COMPETE and by FCT Contract grant number: PEst-C/MAR/LA0017/2013 Published online 00 Month 2014 in Wiley Online Library (wileyonlinelibrary.com). DOI: 10.1002/tox.22070

Among contaminants, inorganic compounds such as metals and metalloids, even at low concentration, may contribute to the estuaries deterioration affecting the organisms inhabiting these areas (Coelho et al., 2006; El-Nemr et al., 2012), particularly benthic macrofauna. Bivalves, which are mostly filter-feeders contacting directly with contaminated water and sediment, are greatly affected by these contaminants (Pellerin and Amiard, 2009; Al-Subiai et al., 2011; Moschino et al., 2012; Jebali et al., 2014). Furthermore, metals and metalloids can be accumulated in lower trophic levels and biomagnified along the food chains (Figueira et al., 2011). When accumulated, elements present a subcellular distribution within an organism (Wallace et al., 2003). The elements accumulated in the soluble fraction are present in the cytosol along with proteins for detoxification (metallothionein-like proteins), and it has been shown that elements present in this fraction are more available for assimilation by predators. The absorption of elements present in the insoluble fraction (metal-rich granules, organelles, and cellular

C 2014 Wiley Periodicals, Inc. V

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debris) is mostly dependent on the species digestive capacity (Wallace and Luoma, 2003; Geffard et al., 2010). To assess the impacts of inorganic contamination in marine bivalves, biochemical alterations, namely associated with oxidative stress, have proved to be of prime relevance (Valavanidis et al., 2006; Alves de Almeida et al., 2007; Freitas et al., 2012a; Moschino et al., 2012). Elements, including lead, cadmium, arsenic, and mercury, can induce oxidative stress in marine bivalves through the production of reactive oxygen species (ROS), namely the superoxide anion (O2 2 ), the hydrogen peroxide (H2O2) and the hydroxyl radical (HO), by mitochondria (Ercal et al., 2001; Valavanidis et al., 2006; Figueira et al., 2012; Regoli and Giuliani, 2014). Most of the ROS generated by cellular metabolism have a special affinity for lipids, proteins and nucleic acids (DNA), causing damage to the cells (Shinde et al., 2012). However, cells also have antioxidant enzymes such as superoxide dismutase (SOD), catalase (CAT), and glutathione Stransferases (GSTs) that can intercept ROS, protecting molecular targets against oxidative injuries (Ramos-Gomez et al., 2011) and increasing the tolerance of organisms in polluted environments (Livingstone, 2001; Freitas et al., 2012b). The antioxidant enzymes SOD and CAT convert O2 2 into H2O2, and H2O2 into molecular oxygen and water, ˇ uracˇkova, respectively (Alves de Almeida et al., 2007; D 2010). The reduced glutathione (GSH) is also an antioxidant useful tool to regulate intracellular homeostasis cycling. Hydrogen peroxide is the substrate for glutathione peroxidase (GPx), using GSH as electron donor to catalyze the reduction of H2O2 into H2O. Some enzymes of the GSTs group reduce lipid hydroperoxides to alcohols with the simultaneous oxidation of GSH to oxidized glutathione (GSSG). Metallothioneins have also received attention as a biomarker of inorganic contamination due to the involvement in protection against metals, oxidant damage, metabolic regulation, and detoxification of toxic metals (PaulPont et al., 2010; Serafim and Bebianno, 2010; Freitas et al., 2012b). In the last years, a large number of marine bivalves, such as Venerupis philippinarum (Sfriso et al., 2008; Moschino et al., 2012; Matozzo et al., 2012; Wang et al., 2012), Cerastoderma edule (Bergayou et al., 2009; Freitas et al., 2012b), V. decussata (Cravo et al., 2012; Matozzo et al., 2012; El-Nemr et al., 2012), Pinctada radiata, Saccostrea cucullata, Circentia callipyga, Pinna muricata (Mora et al., 2004), Pinna nobilis (Jebali et al., 2014), Mytilus edulis (Pellerin and Amiard, 2009), Mya arenaria (Pellerin and Amiard, 2009; Al-Subiai et al., 2011), Perna viridis (Jena et al., 2009), and Venerupis corrugata (Fernandez et al., 2013) have been used as indicators for the presence of toxic substances in the marine environment. Bivalves are a good bioindicator species, reflecting changes in the environmental pollutant status (Box et al., 2007), but they have also an important economic relevance as commercial resource for human consumption. The global aquaculture production of mollusks (mostly bivalves) was

Environmental Toxicology DOI 10.1002/tox

75% (13.9 million tonnes) of total aquaculture, representing an economically relevant activity for worldwide population (Figueira and Freitas, 2013). In Portugal, total production of clams was around 5200 tonnes, in 2012 (FAO, 2012), being important for the national economy. Among clams, one of the most commercialized species in Portugal is V. corrugata (formerly known as V. pullastra) (Gmelin, 1791), easily collected and with high nutritional value (Anacleto et al., 2013b). V. corrugata commonly known as the pullet carpet shell, occurs in Spain, Italy, Portugal, and France (FAO, 2014). This species lives buried in sand, gravel or mud bottoms, usually from intertidal areas (Anacleto et al., 2013a; FAO, 2014). The optimal grow salinity ranges between 21 and 28 g/L (Carregosa et al., 2014) and optimal growth temperature is around 20 C (Anacleto et al., 2013a). Despite its relevance as commercial resource for human consumption (Anacleto et al., 2013b;c; FAO, 2014) and wide spatial distribution, little information is available about V. corrugata, namely in what concerns to its tolerance to inorganic contaminants, bioaccumulation patterns, and biochemical responses under environmental conditions. Therefore, there is an urgent need to increase the knowledge on V. corrugata, especially in terms to the use of this species as a bioindicator. Being one of the most important hotspots for biodiversity in western Iberia, the Ria de Aveiro lagoon system is the ideal location for such research (Freitas et al., 2014). Thus, the objectives of the present work were: (a) assess the element accumulation and intracellular partitioning in V. corrugata collected in areas characterized by different metals and As concentrations; (b) compare sediment and organism contamination, evaluating the bioaccumulation factor (BAF) values obtained in different areas; (c) evaluate the biochemical performance of V. corrugata from areas with different contamination levels; (d) determine relevant biomarkers that could be related to sediment and/or organisms contamination levels.

MATERIAL AND METHODS Field Sampling The area considered for this study was the Ria de Aveiro, located at the northwest Atlantic coast of Portugal (40 380 N, 8 450 W). This aquatic system presents 45 km long and 10 km wide, being connected to the Atlantic Ocean through an artificial channel. This lagoon covers an area between 66 and 83 km2, at low and high tide (spring tide), respectively, comprising several channels (S. Jacinto, Ilhavo, Mira, Ovar, and Murtosa) and distinct intertidal zones, such as mud flats and salt marshes (Dias et al., 2000). The Ria de Aveiro has an important biodiversity in terms of macrofauna (Rodrigues et al., 2011). The presence of bivalves in this ecosystem is not only ecologically important, but also has an economic, social, and cultural importance (Oliveira et al., 2013).

METALS AND ARSENIC CONCENTRATIONS IN SEDIMENTS AND CLAMS

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Fig. 1. Study area showing the localization of Venerupis corrugata in 5 sampling areas (A–E) at the Ria de Aveiro lagoon.

Sampling Procedure The specimens of V. corrugata were collected in 5 areas along the Ria de Aveiro (Fig. 1), characterized by different metals and As contamination levels. At each area, three sites were selected and at each site all V. corrugata specimens present in a square of 50 3 50 cm2 were collected, weighed, and measured. The clams analyzed had average values of 10.95 6 4.20 g for weight, 3.64 6 0.69 cm for length, and 2.61 6 0.51 cm for width. The clam soft tissues of 5 replicates per site were used for elements quantification [chromium (Cr), nickel (Ni), copper (Cu), lead (Pb), cadmium (Cd), mercury (Hg), and arsenic (As)] and also for biochemical analysis. At each sampling site, sediment samples were collected for total organic mat-

ter (TOM) content determination, sediment grain size analysis, and the quantification of elements concentration. In the field, at each area, the environmental variables pH, salinity, and temperature were measured with specific probes. After sampling, V. corrugata individuals and sediments were transported on ice (0 C) to the laboratory and preserved at 220 C until analysis.

Laboratory Analysis Sediment Grain Size and Organic Matter Content Determination Sediment grain-size was analyzed by wet and dry sieving following the procedure described by Quintino et al. (1989).


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The median value, P50, was calculated and expressed in phi (/) units, corresponding to the diameter that has half the grains finer and half coarser (dry weight). The median and the percent content of fines were used to classify the sediment, according to the Wentworth scale: very fine sand (median between 3–4/); fine sand (2–3/); medium sand (1–2/); coarse sand (0–1/); very coarse sand (21 to 0/). The silty and clay fraction was obtained by wet sieving through a 0.063 mm mesh screen and classified with “clean“, “silty,” or “very silty” according to fraction ranges (0–5%, 5–25%, and 25–50%, respectively) of the total sediment, dry weight (Doeglas, 1968; Larsonneur, 1977). Samples with more than 50% fines content were classified as mud. The TOM content was determined according to Byers et al. (1978), being measured as the percentage of weight loss in 1 g of dried sediment, after combustion at 450 C (with minimal risk of volatizing inorganic carbon) during 5 h.

Metals and As Levels Quantification The concentration of seven elements, Cr, Ni, Cu, Pb, Cd, Hg, and As, was measured in organisms (soluble and insoluble fractions) and sediments, following the methodology described by Freitas et al. (2012b;c). The soluble fraction was defined by Wallace and Luoma (2003) as the elements soluble in cytosol, while the insoluble fraction contains the elements in the organelles (MRG, metal-rich granules) and cellular debris. Element quantification was done by ICP-MS (Inductively Coupled Plasma-Mass Spectrometry), in a certified laboratory at the University of Aveiro. Regarding the quality controls, the calibration of the apparatus was made with IV standards and they were verified with standard reference material (National Institute of Standards and Technology, NIST SRM 1643e). During element analysis, the accuracy observed ranged between 90 and 110% (information given by the laboratory). All samples below this accuracy level were rejected and the analysis repeated. Clam determinations were performed by three replicates. The concentration of elements was expressed in microgram per gram fresh weight (FW) and the percentage of soluble and insoluble fraction was calculated. Following McGeer et al. (2003) works, the Bioaccumulation Factor (BAF) was determined dividing the total concentration of a given element in the soft tissue of an organism by the concentration of that element in the sediment, under environmental conditions where organisms are influenced by water and dietary sources.

Biochemical Analysis The specimens collected at each sampling site were weighted, measured, and pulverized with liquid nitrogen. For biochemical analysis 0.5 g of soft tissue was used. Samples extraction was performed with the specific buffer for


each biochemical parameter, centrifuged (10,000g) for 15 min at 4 C. Supernatants were stored at 220 C or used immediately to measure the indicators of cellular damage (total soluble protein; Lipid Peroxidation, LPO; Total glutathione, GSHt), antioxidant and biotransformation enzymes (Catalase, CAT; Superoxide dismutase, SOD; Glutathione S-transferases, GSTs), and metallothioneins (MTs). Indicators of Cellular Damage. Total soluble protein content was determined according to Robinson and Hogden (1940), following the Biuret method, and using bovine serum albumin (BSA) as standard (0–40 mg/mL). The incubation was during 10 min, at 30 C. At the end of this period absorbance was read at 540 nm. The results were expressed in mg per g of fresh weight (FW). LPO quantification was based on the reaction of LPO byproducts, such as malondialdehyde (MDA), with 2thiobarbituric acid (TBA), forming TBARS, according to the protocol described by Buege and Aust (1978). The amount of TBARS, namely MDA, was quantified spectrophotometrically and measured at a wavelength of 532 nm. The calculation of MDA concentration was made using its extinction coefficient (1.56 3 105 M21 cm2). The results were expressed as nmol of MDA equivalents per g of FW. GSHt was quantified according to the DTNB-glutathione reductase (GR) recycling assay (Anderson, 1985). The standard curve was determined using GSSG standards (0–500 mmol/L). The samples and standards were incubated for 5 min at room temperature. Absorbance was measured at 412 nm and GSHt was expressed as lmole per g of FW. Antioxidant (SOD and CAT) and Biotransformation (GSTs) Enzymes. For SOD quantification, the method described by of Beauchamp and Fridovich (1971) was followed, with slight modifications (Freitas et al., 2012a;b). The standard curve was performed with SOD standards (0.25–60 U/mL). SOD activity was measured spectrophotometrically at 560 nm and expressed in U per g of FW. One unit of enzyme (U) corresponds to a reduction of 50% of nitroblue tetrazolium (NBT). CAT activity was determined by the reaction of this enzyme with methanol in the presence of H2O2 (Johansson and Borg, 1988), with some modification according Freitas et al. (2012b). The standard curve was determined using formaldehyde standards (0–150 mM). The incubation was during 20 min in a shaker, at room temperature. The formaldehyde formation with purpald was measured at 540 nm. One unit of enzyme (U) is defined as the amount of enzyme that caused the formation of 1.0 nmol formaldehyde, per min, under the assay conditions. The results were expressed as U of g FW. For the GSTs activity, cell extracts were homogenized in 50 mM phosphate buffer (pH 5 7.0) containing Triton X-100 at 0.1%. The activity of GSTs was determined following the method described by Habig et al. (1974). These


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TABLE I. Sediment physico-chemical parameters, along the 5 sampling areas: temperature ( C); pH; salinity (g/L); sediment type, gravel, sand, and fines content (percentage of total sediment dry weight); the total organic matter (TOM, %) mean values (6standard deviation; n 5 5) Areas

Temperature

pH

Salinity

Main Sediment Type

Median

Gravel

Sand

Fines

TOM

A B C

19.3 20.1 21.6

7.57 7.71 7.04

36 40 35

1.71 6 0.46 2.00 6 0.64 2.30 6 0.53

0.05 6 0.07 0.43 6 0.40 0.00 6 0.00

96.68 6 4.37 86.14 6 13.58 73.81 6 23.63

0.42 6 0.60 19.77 6 10.38 32.72 6 13.44

0.47 6 0.26 0.96 6 0.78 0.58 6 0.21

D E

19.9 21.6

7.21 7.35

33 33

Clean medium sand Silty medium sand Very silty medium sand Mud Silty fine sand

>4 2.10 6 0.66

0.74 6 0.64 6.01 6 4.32

26.18 6 18.90 80.00 6 4.91

73.35 6 17.12 15.40 6 4.17

6.15 6 1.10 3.03 6 1.26

enzymes catalyze the conjugation of the substrate 1-chloro2,4-dinitrobenzene (CDNB) with glutathione, forming a thioester. This formation was followed by the absorbance increment at 340 nm, intervals of 10 s during 5 min. For the enzyme activity quantification it was selected a time interval (5 min) during which the activity was linear. The activity of GSTs was determined using extinction coefficient of 9.6 mM21 cm21 for CDNB. One unit of enzyme (U) corresponds to the amount of enzyme that caused the formation of 1 lmol of thioster per min under the assay conditions. Results were expressed as U per g of FW. Metal Sequestration—Metallothioneins. Proteins separation was done by SDS–PAGE, carried out in 4–20% of polyacrylamide (Mini-PROTEAN TGX—Bio-Rad) according the procedure described by Laemmli (1970). Gels were stained with Coomassie Brilliant Blue R-250 and screened in a Densitometer apparatus (Bio-Rad—Model GS 710). Molecular mass and relative protein amount corresponding to each band were compared with a protein standard (NZY Color Protein Marker II—NZY Tech Genes and Enzymes). Concentration of proteins was calculated using Quantity One program software (Bio-Rad). After separation of proteins each band was cut, and the extraction was done according to Milnerowicz and Bizon (2010). Confirmation of MTs was done through quantification of thiols groups, according to Moron et al. (1979).

Multivariate Analyses Data on element concentration in sediment and organisms as well as biochemical parameters were submitted to hypothesis test through permutation multivariate analysis of variance (PERMANOVA) from PRIMERv6, following the calculation of Euclidean distance matrices among samples. The t-statistic in the pair-wise comparisons was evaluated and values lower than 0.05 revealed significant differences. The significant differences among conditions were identified in Figures with distinct letters. The null hypotheses tested were: (i) for total element concentration in the sediments: no significant differences exist among areas; (ii) for organism’s contamination (considering total contamination, or different element concentration or elements concentration in the solu-

ble and insoluble fractions): no significant differences exist among organisms from different areas; (iii) for each biochemical parameter: no significant differences exist among organisms from different areas. The Spearman correlations were calculated between total element concentration found in the sediment, total element concentration found in the soluble and insoluble fractions, total concentration of elements in clams and biochemical parameters. The Spearman correlation value was classified as strong (0.89 > r > 0.70) and very strong (1.00 > r > 0.90).

RESULTS Sediment Grain Size and Organic Matter Content The results of the sediment physico-chemical characteristics are summarized in Table I. In the areas A, B, C, and E the sediment grain size varied from medium to fine sand, with a median value ranging from 1.71 to 2.10, and the percentage of the fines fraction lower than 33%. Area D was classified as mud (median value higher than 4U) with the fines content higher than 73%. Regarding the TOM content (cf. Table I), the highest values were found at areas D and E (6.15% and 3.03%, respectively), comparing to areas A, B, and C (1.62, 0.59, and 0.56%, respectively).

Metals and As Levels in Sediments and Clams The total element concentration in sediments revealed significantly lower values in the areas located at the center of the lagoon (areas A, B, and C) than in sediments from areas located at the Ilhavo channel (areas D and E), with an increasing contamination gradient from area A to area E [Fig. 2(a) and Table II]. In fact, area A presented significantly lower element concentration than the remaining areas, while areas D and E showed significantly higher contamination values when comparing with the remaining areas. The results obtained also showed that, except in the area E, Cr was the most abundant element in all areas, followed by Pb (cf. Table II). The results further revealed that all


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Fig. 2. Total element concentration in sediments (a) and in Venerupis corrugata (b) mean values (6standard deviation; n 5 6), along the 5 sampling areas (A, B, C, D, E). Significant differences (p 0.05) among areas, are presented with letters (a–c).

elements presented the highest concentration at areas D and E (cf. Table II). The results obtained for the total element concentration found in V. corrugata soluble and insoluble fractions are

present in Figure 2(b). Observing the total concentration of elements in the soluble fraction, no significant differences were noticed among areas, except for area B that presented significantly lower concentration comparing to the

TABLE II. Elements concentration (lg/g dry weight) in sediments, along the sampling areas Areas A B C D E

Cr

Cu

As

Cd

Pb

Hg

Ni

Total

1.22 6 0.07 3.47 6 2.77 5.59 6 2.20 16.95 6 3.21 8.57 6 3.71

0.68 6 0.39 1.89 6 1.61 1.43 6 0.65 4.46 6 1.61 26.22 6 8.99

0.82 6 0.16 1.44 6 1.43 2.27 6 0.91 4.21 6 1.98 4.78 6 1.89

0.02 6 0.00 0.05 6 0.02 0.07 6 0.02 0.20 6 0.05 0.17 6 0.05

0.93 6 0.02 2.18 6 2.24 2.83 6 0.82 5.33 6 1.93 5.19 6 0.00

0.01 6 0.01 0.01 6 0.01 0.03 6 0.01 n.d. 0.06 6 0.02

0.68 6 0.13 1.28 6 1.63 2.19 6 0.82 8.61 6 2.72 4.76 6 2.32

4.37 6 0.11 10.33 6 1.39 14.42 6 0.74 39.78 6 1.64 49.73 6 2.83

The most abundant element in each of the study areas is highlighted in gray.


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TABLE III. Elements (Cr, Cu, Cd, Pb, Hg, Ni, As) concentration (mg/g fresh weight in the soluble and insoluble fractions) and total concentration (mg/g fresh weight) in Venerupis corrugata Fractions A Soluble Insoluble Total B Soluble Insoluble Total C Soluble Insoluble Total D Soluble Insoluble Total E Soluble Insoluble Total

Cr

Cu a

0.15 6 0.15 0.35 6 0.23a 0.50 6 0.15a,b 0.13 6 0.04a 0.66 6 0.14a,b 0.78 6 0.11a,c 0.24 6 0.14a 1,07 6 0,11c 1.31 6 0.24c 0.11 6 0.01a 0.39 6 0.12a 0.50 6 0.12b,d 0.11 6 0.03a 0.90 6 0.13c,b 1.01 6 0.15c,d

As a

0.46 6 0.10 0.43 6 0.38a,b 0.89 6 0.29a,b 0.00 6 0.00a,b 0.68 6 0.30a,b 0.68 6 0.30a 0.85 6 0.50b 0,82 6 0,07a,b 1.67 6 0.48b 0.11 6 0.20a,b 0.62 6 0.11a 0.73 6 0.31a,b 0.34 6 0.30a,b 0.94 6 0.13b 1.28 6 0.48a,b

Cd a,b

2.20 6 0.67 2.22 6 0.27a 4.42 6 0.81a,b 0.94 6 0.22c 1.72 6 1.47a,c 2.66 6 1.67a 1.23 6 0.37c,b 2,08 6 0,76a 3.31 6 0.83a 4.15 6 0.88b 3.93 6 0.82b,c 8.08 6 1.51b 2.47 6 0.28a 3.96 6 0.79b,c 6.43 6 1.02b

Pb a

0.06 6 0.02 0.06 6 0.05a 0.12 6 0.07a 0.04 6 0.01a 0.08 6 0.02a 0.13 6 0.02a 0.05 6 0.01a 0,08 6 0,02a 0.13 6 0.01a 0.04 6 0.01a 0.06 6 0.02a 0.10 6 0.03a 0.07 6 0.02a 0.08 6 0.02a 0.15 6 0.04a

Hg a,b

0.08 6 0.03 0.27 6 0.16a 0.25 6 0.13a 0.01 6 0.02c 0.14 6 0.03a 0.15 6 0.05a 0.05 6 0.03b,c 0,15 6 0,01a 0.15 6 0.09a 0.06 6 0.01d 0.23 6 0.09a 0.29 6 0.10a 0.06 6 0.02b 0.35 6 0.21a 0.40 6 0.21a

Ni a

Total a,b

0.35 6 0.19 3.30 6 0.88a 0.44 6 0.21a 3.69 6 1.24a a 0.75 6 0.15 6.98 6 1.24a a 0.05 6 0.00 1.16 6 0.21b a,c 0.75 6 0.37 4.03 6 5.20a,b 0.78 6 0.15a 5.19 6 2.19a a,b 0.36 6 0.19 2.77 6 0.95a a 1,02 6 0,38 5.17 6 1.00b a,b 1.37 6 0.52 7.95 6 1.74a 0.34 6 0.08b 4.81 6 1.04a a 0.74 6 0.07 5.97 6 0.77c a,b 1.08 6 0.12 10.78 6 1.65b b 0.24 6 0.05 3.29 6 0.32a 0.96 6 0.06b,c 7.19 6 0.88c 1.20 6 0.07b 10.49 6 0.72b

0.00 6 0.00 0.00 6 0.00a 0.00 6 0.00a 0.00 6 0.00a 0.01 6 0.01a 0.01 6 0.01a 0.00 6 0.00a 0,03 6 0,01a 0.01 6 0.02a 0.00 6 0.00a 0.00 6 0.00a 0.00 6 0.00a 0.00 6 0.00a 0.01 6 0.01a 0.01 6 0.01a

Significant differences (p 0.05) among areas for each element concentration in the soluble and insoluble fractions are presented with letters (a–d). The most abundant element, in clams from each of the areas, is highlighted in gray.

remaining areas [cf. Fig. 2(b)]. When analyzing the total concentration of elements present in the insoluble fraction, it is possible to observe that areas with the highest sediment contamination [areas D and E, cf. Fig. 2(a)] presented the highest concentration of elements in this fraction, with significant differences to the remaining areas [cf. Fig. 2(b)]. Furthermore, area B was the area with the lowest element concentration in the insoluble fraction [cf. Fig. 2(b)] and the concentration of elements presented in this fraction was correlated with contamination present in sediments (0.96). Table III presents the results for each element partitioning revealing that, except for Cd and Hg, most of the elements presented higher concentration in the insoluble fraction in comparison with the soluble fraction (cf. Table III). Analyzing the concentration of each element in each area (cf. Table III) it is possible to observe that As was the most abundant element in clams from all the study areas, followed by Cu. Table III also presents the total element concentration in clams from each area. In areas D and E, the total element concentration in clams was significantly higher than in areas (A, B), with clams from area B showing the lowest total element concentration. When comparing the element concentration in clams with the sediments contamination, the BAF values revealed that higher values were found in specimens from the less contaminated areas (areas A, B, and C; BAF > 0.5) (Table IV). Furthermore, clams collected at area A revealed the highest capacity to bioaccumulate contaminants since the BAF value obtained for this area was higher than 1. Table IV also represents the BAF for each element, revealing that higher values were specially noticed for As and Cd, with the highest values at the less contaminated areas (area A and B). Cr and Pb

were the elements that showed the lowest BAF values (cf. Table IV).

Biochemical Responses Indicators of Cellular Damage The results obtained for protein content, along all areas, is shown in Figure 3. The results showed that clams from areas B and C presented lower protein content than clams collected at areas A, D and E. Nevertheless, regarding the clam protein content, no significant differences were found between areas (cf. Fig. 3). Regarding the LPO [Fig. 4(a)], clams from areas B and C showed the lowest values, while clams collect at area D showed the highest values, with significant differences between area D and areas B and C.

TABLE IV. Bioaccumulation factor (BAF) of Venerupis corrugata, for different elements BAF Area A B C D E

Cr

Cu

As

Cd

Pb

Hg

Ni

Total

0.41 0.23 0.23 0.03 0.12

1.30 0.36 1.17 0.16 0.05

5.41 2.51 1.46 1.92 1.34

5.04 2.62 1.90 0.47 0.89

0.37 0.07 0.07 0.05 0.08

0.00 1.51 0.84

1.15 0.54 0.63 0.13 0.25

1.62 0.59 0.56 0.27 0.21

0.21

Bioaccumulation factor: ratio between total element concentration in the organism/total element concentration in the sediment. Values are the mean of three replicates. Highlighted cells represent BAF > 1.


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TABLE V. Spearman correlation between the biochemical parameters and total element concentration in sediments and the total concentration of elements in the soluble and insoluble fractions

Soluble Insoluble SED

LPO

Protein

SOD

CAT

GST

GSHt

MTs

0.77 0.45 0.37

0.70 0.60 0.51

0.00 0.00 20.01

0.71 0.46 0.37

0.70 0.28 0.28

0.70 0.60 0.70

0.36 0.77 0.86

The highest correlations are highlighted in gray.

Fig. 3. Total protein content, mean values (6standard deviation; n 5 6), in Venerupis corrugata along the 5 sampling areas (A, B, C, D, E). Significant differences (p 0.05) among areas, for V. corrugata, are presented with letters (a).

The GSHt content is shown in Figure 4(b). The results obtained showed that the content of this enzyme was significantly lower in clams from areas B and C than in the areas A, D, and E [cf. Fig. 4(b)]. When correlating LPO and protein content with clams (soluble and insoluble fraction) and sediments contamination, the results revealed a strong correlation between both LPO (0.77) and protein content (0.70) and the element concentration present in the soluble fraction (Table V). The

GSHt values, obtained in all areas, were also correlated (0.70) with elements concentration found in sediment and elements concentration present in soluble fraction, and no correlation was found between GSHt and the sediments contamination (cf. Table V).

Antioxidant (SOD and CAT) and Biotransformation (GSTs) Enzymes The activity of the enzymes SOD, CAT, and GSTs are shown in Figures 5(a–c), respectively. The results obtained showed significantly higher SOD activity in areas B, C, and D when compared with areas A and E. On the other hand, the activity of CAT was higher in areas D and E. Nevertheless, only the activity of CAT recorded at area C was significatively different from the values obtained at areas D and E. The GSTs activity in clams from area B was significantly lower than values obtained at the remaining areas. When compared the activity of the enzymes with the clam’s bioaccumulation (soluble and insoluble fractions) and the contamination of sediments, the results showed a higher correlation between the CAT activity and the concentration of elements found in the soluble fraction (0.71) (cf. Table V). Furthermore, the activity of SOD, CAT, and GSTs showed no correlation with sediments contamination (cf. Table V).

Metallothioneins MTs synthesis was significantly higher in clams from areas C, D, and E than in clams from areas A and B (Fig. 6), with area A presenting the lowest and area E the highest MTs values. When analyzing the correlation between MTs and the elements concentration in sediments and clams, the results obtained showed a strong correlation (0.86) between MTs and both the concentration of elements in the insoluble fraction (0.77) and in the sediments (0.86) (cf. Table V). Fig. 4. Indicators of oxidative stress (LPO, lipid peroxidation (a); GSHt, total glutathione (b), mean values (6standard deviation; n 5 6), in Venerupis corrugata along the 5 sampling areas (A, B, C, D, E). Significant differences (p 0.05) among areas, for V. corrugata, are presented with letters (a–b).


DISCUSSION The present study assessed the relationship between metals and As concentration present in the sediments, V. corrugata


9

(4.37 6 0.39 to 14 6 2 mg/g DW). Nevertheless, the total element concentration that characterized the study areas was lower than values found in different locations (among others, Mora et al., 2004; Cheggour et al., 2005; Martın-Dıaz et al., 2007; Ramos-Gomez et al., 2011; Moschino et al., 2012). In this study, the concentration of all elements in sediments from the study areas were lower than values found by other authors in different ecosystems, namely in Moroccan estuaries (90–155 mg/g DW of Cu, Cd, and Ni; Cheggour et al., 2005), and the Venice Lagoon (67–1630 and 79–252 mg/g DW of Cr, Ni, Cu, Pb, Cd, Hg, and As; Sfriso et al., 2008; Moschino et al., 2012, respectively). Similar findings were also found in Santander Bay (31–209 mg/g DW of Cd, Pb, Cu, Ni, Hg; Ramos-Gomez et al., 2011) and at the Indian River Lagoon (43–113 mg/g of Cr, Ni, Pb, Cd As; Trefy and Trocine, 2011). Furthermore, except for Cu in the most contaminated area, in all areas the elements quantified presented concentrations lower than the corresponding TEL (Macdonald et al., 1996) and similar or even lower than values found in pristine areas (Ahn et al., 1996). The analysis on the total element concentration found in V. corrugata individuals (ranging from 5 to 11 mg/g FW) showed values lower than the ones found in marine bivalves from other areas, namely Gulf and Gulf of Oman (192 mg/g DW; Mora et al., 2004), Northern Adriatic lagoons (32–81 mg/g FW; Sfriso et al., 2008 and 19–64 mg/g DW; Moschino et al. 2012), Indian River Lagoon (103–237 mg/g FW; Trefy et al., 2011). Analyzing the concentration of each element, the present work revealed that V. corrugata presented concentrations lower than the values found in clams from different ecosystems. Previous works in four Moroccan estuaries demonstrated that the clam Scrobicularia plana (da Costa, 1778) accumulated higher concentration of Cd, Cu, and Ni (39.99–59.14 mg/g DW; Cheggour et al., 2005) when compared with V. corrugata from the Ria de Aveiro. Sfriso et al. (2008) also showed that V. philippinarum, from two lagoons of the northern Adriatic Sea, presented higher concentration Fig. 5. Enzymes (SOD, superoxide dismutase; CAT, catalase; GST, glutathione Stransferase) in Venerupis corrugata along the 5 sampling areas (A, B, C, D, E), mean values (6standard deviation; n 5 6). Significant differences (p 0.05) among areas, for Venerupis corrugata, are presented with letters (a–b).

bioaccumulation and the biochemical performance of this native species collected from 5 areas, resembling an increasing sediment contamination gradient in the Ria de Aveiro lagoon (Portugal). To our knowledge, this is the first study relating the biochemical responses of V. corrugata with the species bioaccumulation and the environmental contamination. The results obtained showed that sediments from the areas located at the Ilhavo channel presented higher elements concentration (40 6 10 to 50 6 8 mg/g DW) than the sediments from the areas at the center of the lagoon

Fig. 6. Metalotioninas (MTs) content, mean values (6standard deviation; n 5 6), in Venerupis corrugata along the 5 sampling areas (A, B, C, D, E). Significant differences (p 0.05) among areas, for V. corrugata, are presented with letters (a–b).


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of As, Cd, Cr, Cu, Pb, Ni, and Hg than V. corrugata. Studies by Koch et al. (2007) demonstrated that the clam Mya arenaria presented higher As content (7.0–7.9 mg/g) than clams used in the present study. Valette-Silver et al. (1999) also found higher values of As (23 mg/g) in the oyster Crassostrea virginica. For the Ria de Aveiro previous studies with Cerastoderma edule (Linnaeus, 1758) demonstrated that cockles accumulated similar or higher amount of most of the metals tested here (Cr, Cu, Cd, Pb, Hg, and Ni) but lower amount of As (Figueira et al., 2011), evidencing the difference in As accumulation between the two species. The present study further revealed that As was the most abundant element in clams from all the study areas. This is explained by the fact that As is one of the most abundant elements in the Ria de Aveiro (the present study and Freitas et al., 2014) and V. corrugata presented a high bioaccumulation capacity (BAF > 1) for this element. Sfriso et al. (2008) demonstrated that, when comparing among different elements. As was one of the most abundant in V. philippinarum from Northern Adriatic lagoons. Moschino et al. (2012) also showed that, among different elements, As was the most abundant in V. philippinarum from the Venice lagoon. Concerning clams bioaccumulation, the present work revealed that in areas with higher contamination levels clams showed higher concentration of elements. However, in general, clams collected in areas with less sediment contamination present higher BAF values indicating higher bioaccumulation. In fact, V. corrugata specimens collected at the less contaminated area presented higher bioaccumulation capacity than clams from the remaining areas, which was demonstrated by the BAF higher than 1. These results are in accordance with previous works suggesting that the high availability of metals in sediments from the less contaminated area may be explained by the lowest percentage of TOM (Yu et al., 2012). Works conducted by Zhong and Wang (2009) demonstrated that natural organic matter (NOM), such as humic acid, fulvic acid, and amino acids, has a strong binding affinity for trace metals in the aquatic environment. According to Freitas et al. (2012b), C. edule also accumulated higher elements concentration in areas less contaminated when compared with highly contaminated areas. When analyzing the different elements in clams, it is possible to observe that, in most of the areas, BAF values were higher than 1 for Cu, As, and Cd, revealing that clams tend to bioaccumulate preferentially these elements. Furthermore, for all elements, higher BAF values were found at the less contaminated areas. For C. edule Figueira et al. (2011) reported higher BAF values for all the analyzed elements in the less contaminated area and, except for Al and Cr, BAF values were above 1. Whyte et al. (2009) also found the highest concentration of Cd in mussels from clean sites. Other studies reported similar results with higher Cd concentrations in scallops from clean sites compared to specimens collected in contaminated areas (Uthe and Chou, 1987; Bustamante and Miramand, 2005).


Regarding the element partitioning in clams, the present work revealed that the total element concentration in clams soluble fraction did not follow the sediment contamination and was less variable than the levels found in sediments from the different areas. This fact suggests that, the concentration of metals and As in the clams soluble fraction is not proportional on the contamination levels found in the sediments. Thus, the risk to predatory species, through trophic chain transference, might be similar among the different areas. Furthermore, the results obtained revealed that, for all areas, the elements analyzed were predominantly in the insoluble form, being in the precipitated form or bound to membranes (Wallace and Luoma, 2003; Wallace et al., 2003), with significant differences among the studied areas. This is in agreement with previous studies, which demonstrated that in marine bivalves the insoluble form is predominant (Hamza-Chaffai, et al., 2000; Pellerin and Amiard, 2009; Rainbow and Smith, 2010). The biological response of bivalves species, under different conditions, were already described by several authors (among others, Valavanidis et al., 2006; Alves de Almeida et al., 2007; Box et al., 2007; Bergayou et al., 2009; Jena et al., 2009; AlSubiai et al., 2011; El-Nemr et al., 2012; Freitas et al., 2012b; Wang et al., 2012; Regoli and Giuliani, 2014) but there are no studies assessing the biochemical responses of V. corrugata when under different environmental contamination levels. In the present work, biochemical responses of V. corrugata from different contaminated areas were assessed through proteins and LPO content, total glutathione content, antioxidant enzymes (CAT, SOD, GSTs) and MTs. The results obtained showed a higher LPO and protein content in areas with lower and higher levels of contamination. When comparing the lipid peroxidation and the protein content with total elements present in the soluble fraction results showed a strong correlation. This suggests that, elements present in the cytosol increases the ROS formation, inducing LPO. According to Alves de Almeida et al. (2007) exposing aquatic organisms to metals can increase the production of ROS, being the cell membranes potential targets, thus leading to oxidative degradation of lipids known as lipid peroxidation. In fact, increasing levels of LPO in gills of S. plana sampled from contaminated areas were described by Ahmad et al. (2011). Also Zhang et al. (2010) found higher LPO levels in the bivalve Chlamys farreri (Jones et Preston, 1904) when exposed to environmentally relevant concentrations of lead. According to Hamza-Chaffai et al. (2000) and Cravo et al. (2012), LPO was considered an indicator of environmental stress, employed as a damage biomarker reflecting exposure and toxicity to inorganic contaminants by the clam V. decussata (Linnaeus, 1758). The most abundant thiol compound in cells playing an important protective role against oxidative stress is glutathione (Gibson et al., 2012). In the present work, the results showed a higher GSHt content in areas with higher


contamination. The GSHt increase in the mussels Mytilus edulis (Linnaeus, 1758) and Perna perna (Linnaeus, 1758), subjected to metal contamination, was also reported by AlSubiai et al. (2011) and Drafre et al. (2004), respectively. In agreement with these studies, the mussel Mytella guyanensis (Lamarck, 1819) from polluted mangroves, showed an increase in GSHt content when compared with mussels from a control area (Torres et al., 2002). The enzymes SOD, CAT, and GSTs were reported to be sensitive, reliable and able to represent the earliest signals of environmental disturbance being a protection mechanism against oxidative stress (Matozzo et al., 2012; Wang et al., 2012). In the present study the results obtained showed that, except at the most contaminated area, the SOD activity increased with the increasing sediment contamination. These findings suggest that at low and moderate contamination levels this enzyme was acting as an antioxidant defense but at higher contamination other pathways of detoxification may be activated. Geret et al. (2003) evaluated the influence of different sources of contamination on antioxidant enzymes in the gills and digestive gland of the clam V. decussata, and showed lower activity of mitochondrial SOD and higher activity of mitochondrial CAT in clams from a low contaminated area. However, recent studies demonstrated that the SOD activity increased in the mussel Dreissena polymorpha (Pallas, 1771) and the clam V. philippinarum (Adams and Reeve, 1850), when exposed to metal contamination, acting as a first line of antioxidant enzymatic defense detoxifying ROS (Faria et al., 2009; Wang et al., 2012). After destruction of the superoxide anion radical through generation of hydrogen peroxide by SOD enzyme, CAT is responsible for the breakdown of hydrogen peroxide and subsequently its degradation into water and oxygen (Regoli and Giuliani, 2014). In the present work results showed that the activity of CAT was higher at the less and the most contaminated areas, which could indicate an increase in oxidative stress in organisms collected in these areas, often connected to excessive oxyradical formation during the catabolism of various organic compounds (Fernandez et al., 2010). In fact, at the most contaminated areas high LPO values were recorded. The performance of CAT as an indicator of oxidative stress caused by different levels of pollution was previously studied in the mussel M. galloprovencialis, and the results obtained revealed that the CAT activity was higher in the most contaminated areas (Box et al., 2007). In the present study, the high CAT activity at the lowest contaminated area may be related to the fact that almost 50% of the total element concentration found in clams is allocated to the soluble fraction, which is the fraction that causes toxicity to the organism, being evidenced by the high LPO present in the organisms from this area. In fact, the present study showed a strong correlation between the CAT activity and the soluble fraction for all areas, suggesting a higher decomposition of H2O2 present in cytosol by this enzyme. Accord-

11

ing to Boveris and Cadenas (2000), mitochondria are a major source of H2O2 through SOD enzyme. Thus, these organelles may increase the cytosolic steady-state level of oxidants. The H2O2 produced in mitochondria can be transferred through diffusion to cytosolic space and/or for mitochondrial matrix and there converted by cytosolic CAT. In fact, Geret et al. (2003) evaluated the influence of different sources of contamination on antioxidant enzymes, in V. decussata, and reported that cytosolic CAT activity plays an as important role of detoxification in the presence of organic xenobiotics. Glutathione S-transferases (GSTs) represent a major group of detoxification isoenzymes (Fernandez et al., 2010). Our results revealed that, in V. corrugata, the GSTs levels were higher in the most and in the less contaminated areas. The higher concentration of metals found in clams from the less contaminated area suggested the unbalance in ROS status and an increase in antioxidant enzymes against oxidative damage. According to Wang et al. (2012), the higher GSTs levels could be due to detoxification processes or antioxidant action against metals, since GSTs provide a first defense mechanism though elimination of reactive compounds by forming conjugates with glutathione. Works by Fernandez et al. (2010), demonstrated that in the mussels M. galloprovincialis and Perna viridis collected in highly metal-polluted sites, the levels of GSTs were significantly higher than the values found at the lower or non-polluted sites. In agreement, Jena et al. (2009) demonstrated that natural population of P. viridis showed high GSTs activity under environmental pollution. Zhang et al., (2010) evaluated the biomarker responses in the bivalve C. farreri of the environmentally relevant concentrations of lead, mercury and copper, which increased GSTs activity when compared with control for Pb and Hg. Metallothioneins (MTs) are metal chelating proteins protecting cells from oxidative stress (Valavanidis et al., 2006; Freitas et al., 2012b). Regarding MTs content, the present work showed an increase in MTs concentrations present in bivalves from areas with higher contamination, suggesting a higher capacity to regulate metal concentration by V. corrugata, providing an important organism resistance against accumulated metals and As. In fact, the increase of MTs was strongly correlated with the increase of the contamination present in sediments, supporting the capacity of V. corrugata to regulate metal concentration present in cells. Bivalves such as M. galloprovincialis (Serafim et al., 2011), V. philippinarum (Martın-Dıaz et al., 2007) and Corbicula fluminea (Baudrimont et al., 1997), were exposed to metal contamination and the results showed an increase of MTs levels with the increase of metals, demonstrating bind ability of metals to MTs. Indeed, this is in agreement with Serafim and Bebianno (2010) findings that revealed that MTs levels were proportional to the increase of Cd and Cu concentrations in V. decussata tissues, increasing during the exposure time and decreasing during the depuration time. Geret et al. (2003) results also showed an


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increase in MTs levels present in the gills and digestive gland of the clam V. decussata from areas influenced by different sources of contamination when compared with control area.

CONCLUSION In conclusion, the clam V. corrugata was collected in 5 areas, characterized by an increasing concentration in metals and As. The results obtained showed that the organism’s bioaccumulation pattern was not in agreement with sediments contamination level. In fact, the bioaccumulation factor calculated for V. corrugata revealed higher values at the less contaminated areas, being higher than 1 in the less contaminated area (area A), indicating the higher element concentration in clams than in the sediments from this area. The results obtained further revealed the higher ability of V. corrugata to accumulate As (BAF > 1) independently on sediment and clams total element concentration. Regarding elements partitioning, in general, the total elements concentration in the soluble fraction showed no significatively differences among the studied areas. In most of the areas and for most of the elements, higher concentrations were found in the insoluble fraction, with most of the elements being precipitated or bound to membranes. The biochemical biomarkers, protein content, LPO content, GSHt content, and CAT activity, were higher in areas with higher sediment contamination (areas D and E) and in the less contaminated area (area A), with strong correlation with total elements present in the soluble fraction. These results suggest that cell membrane was subjected to the attack of ROS and consequently to the increase of oxidative degradation of lipids (LPO) in cytosol. The increase in CAT and GST activity, and in the GSHt content were indicators of oxidative stress caused by exposure to metals and As and consequently this increase induce the defense against ROS. The synthesis of MTs was higher in clams from contaminated areas, suggesting the elimination of metals through these MTs. The present work demonstrated that LPO, CAT, GSHt, and MTs present an important role in maintaining cellular metabolism homeostasis and protecting V. corrugata from environmental pollution and could be used as good biomarkers of metal and As contamination in environmental studies, even at low contaminated areas.

REFERENCES Ahmad I, Mohmood I, Mieiro CL, Coelho JP, Pacheco M. Santos MA, Duarte AC, Pereira E. 2011. Lipid peroxidation versus antioxidant modulation in the bivalve Scrobicularia plana in response to environmental mercuryeorgan specificities and age effect. Aquat Toxicol 103:150–158. Ahn I-Y, Lee SH, Kim KT, Shim JH, Kim D-Y. 1996. Baseline heavy metal concentrations in the Antarctic clam, Laternula


elliptica in Maxwell Bay, King George Island, Antarctica. Mar Pollut Bull 32:592–598. Al-Subiai SN, Moody AJ, Mustafa SA, Jha NA. 2011. A multiple biomarker approach to investigate the effects of copper on the marine bivalve mollusk, Mytilus edulis. Ecotox Environ Safe 74:1913–1920. Alves de Almeida E, Bainy ACD, Loureiro APM, Martinez GR, Miyamoto S, Onuki J, Barbosa LF, Garcia CCM, Prado FM, Ronsein GE, Sigolo CA, Brochini CB, Martins AMG, Medeiros MHG, Mascio P. 2007. Oxidative stress in Perna perna and other bivalves as indicators of environmental stress in the Brazilian marine environment: Antioxidants, lipid peroxidation and DNA damage. Comp Biochem Physiol: Part A 146:588–600. Anacleto P, Maulvault AL, Barrento S, Mendes R, Nunes, ML, Rosa R, Marques A. 2013a. Physiological responses to depuration and transport of native and exotic clams at different temperatures. Aquaculture 408–409:136–146. Anacleto P, Pedro S, Nunes ML, Rosa R, Marques A. 2013b. Microbiological composition of native and exotic clams from Tagus estuary: Effect of season and environmental parameters. Mar Pollut Bull 74:116–124. Anacleto P, Maulvault AL, Chaguri M, Pedro P, Nunes ML, Rosa R, Marques A. 2013c. Microbiological responses to depuration and transport of native and exotic clams at optimal and stressful temperatures. Food Microbiol 36:365–373. Anderson ME. 1985. Determination of glutathione and glutathione disulfide in biological samples. In: Meister A, editor. Methods in Enzymology, Vol. 113. Orlando, FL: Academic Press. pp 548–555. Baudrimont M, Metivaud J, Maury-Brachet R, Ribeyre F, Boudou A. 1997. Bioaccumulation and metallothionein response in the asiatic clam (Corbicula fluminea) after experimental exposure to cadmium and inorganic mercury. Environ Toxicol Chem 16: 2096–2105. Beauchamp C, Fridovich I. 1971. Superoxide dismutase: improved assays and an assay applicable to acrylamide gels. Anal. Biochem 44:276–287. Bergayou H, Mouneyrac C, Pellerin J, Moukrim A. 2009. Oxidative stress responses in bivalves (Scrobicularia plana, Cerastoderma edule) from the Oued Souss estuary (Morocco). Ecotox Environ Safe 72:765–769. Boveris A, Cadenas E. 2000. Mitochondrial production of hydrogen peroxide regulation by nitric oxide and the role of ubisemiquinone. Crit Rev IUBMB Life 50:245–250. Box A, Sureda A, Galgani F, Pons A, Deudero S. 2007. Assessment of pollution at Balearic Islands applying oxidative stress biomarkers in the mussel Mytilus galloprovincialis. Comp Biochem Physiol Part C 146:531–539. Buege JA, Aust SD. 1978. Microsomal lipid peroxidation methods. Enzymology 52:302–310. Bustamante P, Miramand P. 2005. Subcellular and body distributions of 17 trace elements in the variegated scallop Chlamys varia from the French coast of the Bay of Biscay. Sci Total Environ 337:59–73.


Byers C, Mills EL, Steward PL. 1978. A comparison of methods of determining organic carbon in marine sediments with suggestions for a standard method. Hydrobiologia 58:43–47. Carregosa V, Velez C, Soares AMVM, Figueira E, Freitas R. 2014. Physiological and biochemical responses of the three Veneridae clams exposed to salinity changes. Comp Biochem Phys: Part B 177–178:1–9. Cheggour M, Chafik A, Fisher NS, Benbrahim S. 2005. Metal concentration in sediments and clams in four Moroccan estuaries. Mar Environ Res 59:119–137. Coelho JP, Rosa M, Pereira E, Duarte A, Pardal MA. 2006. Pattern and annual rates of Scrobicularia plana mercury bioaccumulation in a human induced mercury gradient (Ria de Aveiro, Portugal). Estuar Coast Shelf 69:629–635. Cravo A, Pereira C, Gomes T, Cardoso C, Serafim A, Almeida C, Rocha T, Lopes B, Company R, Medeiros A, Norberto R, Pereira R, Araujo O, Bebiano MJ. 2012. A multibiomarker approach in the clam Ruditapes decussatus to assess the impact of pollution in the Ria Formosa Lagoon, South Coast of Portugal. Mar Environ Res 75:23–34. Dafre AL, Medeiros ID, Muller IC, Ventura EC, Bainy ACD. 2004. Antioxidant enzymes and thiol/disulfide status in the digestive gland of the brown mussel Perna Perna exposed to lead and paraquat. Chem-Biol Interact 149:97–105. Dias JM, Lopes JF, Dekeyser I. 2000. Tidal propagation in the Ria lagoon. Portugal. Physand Chem Earth PT B 25:369–374. Doeglas DJ. 1968. Grain size indices, classification and environment. Sedimentology 10:8–82. ˇ Duracˇkova Z. 2010. Some current insights into oxidative stress review. Physiol Res 59:459–469. El-Nemr A, Khaled A, Moneer AA, El-Sikaily A. 2012. Risk probability due to heavy metals in bivalve from Egyptian Mediterranean coast. Egypt J Aquatic Res 38:67–75. Ercal N, Gurer-Orhan H, Aykin-Burns N. 2001. Toxic metals and oxidarive stress Part I: Mechanisms involved in metal induced oxidative damage. Curr Top Med Chem 1:529–539.

13

Figueira E, Lima A, Branco D, Quintino V, Rodrigues AM, Freitas R. 2011. Health concerns of consuming cockles (Cerastoderma edule L.) from a low contaminated coastal system. Environ Int 37:965–972. Figueira E, Cardoso P, Freitas F. 2012. Ruditapes decussatus and Ruditapes philippinarum exposed to cadmium: Toxicological effects and bioaccumulation patterns. Comp Biochem Physiol Part C 156:80–86. Freitas R, Pinto LR, Sampaio M, Costa S, Silva M, Rodrigues AM, Quintino V, Figueira, E. 2012a. Effects of depuration on the element concentration in bivalves: Comparison between sympatric Ruditapes decussatus and Ruditapes philippinarum. Estuar Coast Shelf S 10:43–53. Freitas R, Costa E, Velez C, Santos J, Lima A, Oliveira C, Rodrigues AM, Quintino V, Figueira E. 2012b. Looking for suitable biomarkers in benthic macroinvertebrates inhabiting coastal areas with low metal contamination. Ecotox Environ Safe 75:109–118. Freitas R, Pires A, Quintino V, Rodrigues AM, Figueira E. 2012c. Subcellular partitioning of elements and availability for trophic transfer: comparison between the Bivalve Cerastoderma edule and the Polychaete Diopatra neapolitana. Estuar Coast Shelf S 99:21–30. Freitas R, Martins R, Campino B, Figueira E, Soares AMVM, Montaudouin X. 2014. Trematode communities in cockles (Cerastoderma edule) of the Ria de Aveiro (Portugal): Influence of inorganic contamination. Mar Pollut Bull 82:117–126. Geffard A, Sartelet H, Garric J, Biagianti-Risbourg S, Delahaut D, Geffard O. 2010. Subcellular compartmentalization of cadmium, nickel, and lead in Gammarus fossarum: Comparison of methods. Chemosphere 78:822–829. Geret F, Serafim A, Bebianno MJ. 2003. Antioxidant enzyme activities, metallothioneins and lipid peroxidation as biomarkers in Ruditapes decussatus? Ecotoxicology 12:417–426.

FAO. 2012. http://www.fao.org/fishery/statistics/en.

Gibson SA, Kora Z, Shelton RC. 2012. Oxidative stress and glutathione response in tissue cultures from persons with major depression. J Psychiat Res 46:1326–1332.

FAO. 2014. Species Fact Sheets: Venerupis pullastra (Adams and Reeve, 1850). (Retrieved June, 25, 2014 from http://www.fao. org/fishery/culturedspecies/Venerupis_pullastra/en).

Habig WH, Pabst MJ, Jakoby WB. 1974. Glutathione Stransferases. The first enzymatic step in mercapturic acid formation. J Biol Chem 249:7130–7139.

Faria M, Carrasco L, Diez S, Riva MC, Bayona JM, Barata C. 2009. Multi-biomarker responses in the freshwater mussel Dreissena polymorpha exposed to polychlorobiphenyls and metals. Comp Biochem Phys Part C 149:281–288.

Hamza-Chaffai A, Amiard JC, Pellerin J, Joux L, Berthet B. 2000. The potential use of metallothionein in the clam Ruditapes decussatus as a biomarker of in situ metal exposure. Comp Biochem Physiol C 127:185–197.

Fernandez B, Campillo JA, Martınez-Gomez C, Benedicto J. 2010. Antioxidant responses in gills of mussel (Mytilus galloprovincialis) as biomarkers of environmental stress along the Spanish Mediterranean coast. Aquatic Toxicol 99:86–197.

Jebali J, Chouba L, Banni M, Boussetta H. 2014. Comparative study of the bioaccumulation and elimination of trace metals (Cd,Pb, Zn, Mn and Fe) in the digestive gland, gills and muscle of bivalve Pinna nobilis during a field transplant experiment. J Trace Elem Med Biol 28:212–217.

Fernandez N, Fernandez-Boan M, Verısimo P, Freire J. 2013. Assessing the spatial variability, level and source of organic chemical contaminants in bivalve fishing grounds on the Galician coast (NW Spain). Mar Pollut Bull 74:291–301. Figueira E, Freitas R. 2013. Consumption of Ruditapes philippinarum and Ruditapes decussatus: Comparison of element accumulation and health risk. Environ Sci Pollut R 20:5682–5691.

Jena KB, Verlecar XN, Chainy GBN. 2009. Application of oxidative stress indices in natural populations of Perna viridis as biomarker of environmental pollution. Mar Pollut Bull 58:107–113. Johansson LH, Borg LAH. 1988. A spectrophotometric method for determination of catalase activity in small tissue samples. Anal Biochem 174:331–336.


14

VELEZ ET AL

Koch I, McPherson K, Smith P, Easton L, Doe KG, Reimer KJ. 2007. Arsenic bioaccessibility and speciation in clams and seaweed from a contaminated marine environment. Mar Pollut Bull 54:586–594. Laemmli UK. 1970. Cleavage of structural proteins during the assembly of the head of bacteriophage T4. Nature 227:680– 685. Larsonneur C. 1977. La cartographie des depots meubles sur le plateau continental franc¸ais: methode mise au point et utilisee en Manche. J Rech Oceanogr 2:33–39. Livingstone DR. 2001. Oxidative stress in aquatic organisms in relation to pollution and aquaculture. Revue Med Vet 154:427– 430. Macdonald DD, Carr RS, Calder FD. 1996. Development and evaluation of sediment quality guidelines for Florida coastal waters. Ecotoxicology 5:253–278. Martın-Dıaz ML, Blasco J, Sales D, DelValls TA. 2007. Biomarkers study for sediment quality assessment in Spanish ports using the crab carcinus maenas and the clam ruditapes philippinarum. Arch Environ Contam Toxicol 53:66–76. Matozzo V, Binelli A, Parolini M, Previato M, Masiero L, Finos L, Bressan M, Marin MG. 2012. Biomarker responses in the clam Ruditapes philippinarum and contamination levels in sediments from seaward and landward sites in the Lagoon of Venice. Ecol Indic 19:191–205. McGeer JC, Brix KV, Skeaff JM, DeForest D, Brigham SI, Adams WJ, Green A. 2003. Inverse relationship between bioconcentration factor and exposure concentration for metals: Implications for hazard assessment of metals in the aquatic environment. Environ Toxicol Chem 22:1017–1037. Milnerowicz H, Bizon A. 2010. Determination of metallothionein in biological fluids using enzyme-linked immunoassay with commercial antibody. Acta Biochim Pol 57:99–104.

Quintino V, Rodrigues AM, Gentil F. 1989. Assessment of macro zoobenthic communities in the lagoon of Obidos, western of Portugal. Sci Mar 53:645–654. Rainbow PS, Smith BD. 2010. Trophic transfer of trace metals: Subcellular compartmentalization in bivalve prey and comparative assimilation efficiencies of two invertebrate predators. J Exp Mar Biol Ecol 390:143–148. Ramos-Gomez J, Coz A, Viguri JR, Luque A, Martın-Dıaz MLT, DelValls A. 2011. Biomarker responsiveness in different tissues of caged Ruditapes philippinarum and its use within an integrated sediment quality assessment. Environ Pollut 159:1914– 1922. Regoli F, Giuliani ME. 2014. Oxidative pathways of chemical toxicity and oxidative stress biomarkers in marine organisms. Mar Environ Res 93:106–117. Robinson HW, Hogden CC. 1940. The biuret reaction in the determination of serum protein. J Biol Chem 707–725. Rodrigues AM, Quintino V, Sampaio L, Freitas R, Neves R. 2011. Benthic biodiversity patterns in Ria de Aveiro, Western Portugal: Environmental-biological relationships. Estuar Coast Shelf S 95:338–348. Rodrigues M, Oliveira A, Queiroga H, Fortunato AB, Zhang YJ. 2009. Three-dimensional modeling of the lower trophic levels in the Ria de Aveiro (Portugal). Ecol Model 220:1274–1290. Santos L, Vaz L, Gomes NCM, Vaz N, Dias JM, Cunha A, Almeida A. 2014. Impact of freshwater inflow on bacterial abundance and activity in the estuarine system Ria de Aveiro. Estuar Coast Shelf S 138:107–120. Serafim A, Bebianno MJ. 2010. Effect of a polymetallic mixture on metal accumulation and metallothioneins response in the clam Ruditapes decussatus. Aquat Toxicol 99:370–378.

Mora S, Fowler SW, Wyse E, Azemard S. 2004. Distribution of heavy metals in marine bivalves, fish and coastal sediments in the Gulf and Gulf of Oman. Mar Pollut Bull 49:410–424.

Serafim A, Lopes B, Company R, Cravo A, Gomes T,Sousa V, Bebianno MJ. 2011. A multi-biomarker approach in crosstransplanted mussels Mytilus galloprovincialis. Ecotoxicology 20:1959–1974.

Moron MS, Depierre JW, Mannervik B. 1979. Levels of glutathione, glutathione reductase and glutathione S-transferase activities in rat lung and liver. Biochim Biophys Acta 582:67–78.

Sfriso A, Argese E, Bettiol C, Facca C. 2008. Tapes philippinarum seed exposure to metals in polluted areas of the Venice lagoon. Estuar Coast Shelf S 79:581–590.

Moschino V, Delaney E, Ros LD. 2012. Assessing the significance of Ruditapes philippinarum as a sentinel for sediment pollution: bioaccumulation and biomarker responses. Environ Pollut 171:52–60.

Shinde A, Ganu J, Nail P. 2012. Effect of free radicals and antioxidants on oxidative stress: A review. J Dental Allied Sci 1:63– 66.

Oliveira J, Castilho F, Cunha A, Pereira, MJ. 2013. Bivalve harvesting and production in Portugal: An overview. J Shellfish Res 32:911–924. Paul-Pont I, Montaudouin X, Gonzalez P, Soudant P, Baudrimont M. 2010. How life history contributes to stress response in the Manila clam Ruditapes philippinarum. Environ Sci Pollut Res 17: 987–999. Pellerin J, Amiard JC. 2009. Comparative of bioaccumulation of metals and induction of metals and induction of metallothioneins in two marine bivalves (Mytilus edulis and Mya arenaria). Comp Biochem Physiol Part C 150:186–195.


Spada L, Annicchiarico C, Cardellicchio N, Giandomenico S, Leo L. 2012. Mercury and methylmercury concentrations in Mediterranean seafood and surface sediments, intake evaluation and risk for consumers. Int J Hyg Environ Health 215:418–426. Torres MA, Testa CP, Gaspari C, Masutti MB, Panitz CMN, CuriPedrosa R, Almeida EA, Mascio PD, Di Filho W. 2002. Oxidative stress in the mussel Mytella guyanensis from polluted mangroves on Santa Catarina Island, Brazil. Mar Pollut Bull 44: 923–932. Trefy JH, Trocine RP. 2011. Metals in sediments and clams from the Indian River Lagoon, Florida: 2006–7 Versus 1992. Fla Scientist 74:43–62.


Uthe JF, Chou CL. 1987. Cadmium in sea scallop (Plactopecten magellanicus) tissues from clean and contaminated areas. Can J Fish Aquat Sci 44:91–98. Valavanidis A, Vlahogianni T, Dassenakis M, Scoullos M. 2006. Molecular biomarkers of oxidative stress in aquatic organisms in relation to toxic environmental pollutants. Ecotox Environ Safe 64:178–189. Valette-Silver NJ, Riedel GF, Crecelius EA, Windom H, Smith RG, Dolvin SS. 1999. Elevated arsenic concentrations in bivalves from the southeast coasts of the USA. Mar Environ Res 48:311–333. Wallace WG, Luoma SN. 2003. Subcellular compartmentalization of Cd and Zn in two bivalves. II. Significance of trophically available metal (TAM). Mar Ecol Prog Ser 257:125– 137. Wallace WG, Gweon Lee BG, Luoma SN. 2003. Subcellular compartmentalization of Cd and Zn in two bivalves. I. Significance of metal-sensitive fractions (MSF) and biologically detoxified metal (BDM). Mar Ecol Prog Ser 249:183–197.

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Wang Z, Yan C, Vulpe CD, Yan Y, Chi Q. 2012. Incorporation of in situ exposure and biomarkers response in clams Ruditapes philippinarum for assessment of metal pollution in coastal areas from the Maluan Bay of China. Mar Pollut Bull 64:90–98. Whyte A.L.H., Hook GR, Greening GE, Gibbs-Smith E, Gardner JPA. 2009. Human dietary exposure to heavy metals via the consumption of greenshell mussels (Perna canaliculus Gmelin 1791) from the Bay of Islands, northern New Zealand. Sci Total Environ 407:4348–4355. Yu X, Li H, Pan K, Yan Y, Wang, W-X. 2012. Mercury distribution, speciation and bioavailability in sediments from the Pearl River Estuary, Southern China. Mar Pollut Bull 64:1699–1704. Zhang Y, Song J, Yuan H, Xu Y, He Z, Duan L. 2010. Biomarker responses in the bivalve (Chlamys farreri) to exposure of the environmentally relevant concentrations of lead, mercury, copper. Environ Toxicol Phar 30:19–25. Zhong H, Wang W-X. 2009. The role of sorption and bacteria in mercury partitioning and bioavailability in artificial sediments. Environ Pollut 157:981–986.
