Biomarker responses in the bivalve (Chlamys farreri) to exposure of the environmentally relevant concentrations of lead, mercury, copper

Environmental Toxicology and Pharmacology 30 (2010) 19–25

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Biomarker responses in the bivalve (Chlamys farreri) to exposure of the environmentally relevant concentrations of lead, mercury, copper Ying Zhang a,b , Jinming Song a,∗ , Huamao Yuan a , Yayan Xu a,b , Zhipeng He c , Liqin Duan a,b a b c

Key Laboratory of Marine Ecology and Environmental Sciences, Institute of Oceanology, Chinese Academy of Sciences, 7# Nanhai Road, Shandong 266071, PR China Graduate University of Chinese Academy of Sciences, Beijing 100049, PR China Fishery Design Institute of Shandong Province, Ji’nan, Shandong 250013, PR China

a r t i c l e

i n f o

Article history: Received 20 November 2009 Received in revised form 7 March 2010 Accepted 9 March 2010 Available online 16 March 2010 Keywords: Heavy metal Antioxidant enzyme CYP450 Biomarker Bivalve

a b s t r a c t Water samples collected from Bohai Bay were determined to describe the distributions of lead, mercury, and copper in this area, indicating that mean values of the three metals were 1.63 ␮g/L, 4.85 × 10−2 ␮g/L, and 2.68 ␮g/L, respectively. Only lead exceeded the first class limit of seawater quality standard in China. Then, antioxidant enzymes, lipid peroxidation, and metabolic enzymes were investigated in bivalves (Chlamys farreri), exposed to three metals at the environmental concentration levels obtained from our investigation. Significantly reduced SOD, CAT and GPx activities, in lead-exposed group were observed and resulted in obvious lipid peroxidation. In contrast, mercury and copper did not show such clear oxidative stresses. In consistent with the oxidative stress variations, exposed only to lead caused a great inhibition on EROD activity. Multi-biomarker responses in bivalve when exposed to lead at the environmentally relevant concentration in Bohai Bay suggested that lead may possess a potential risk in this area. © 2010 Elsevier B.V. All rights reserved.

1. Introduction Cytochrome P450, the terminal oxidase of monooxygenases, is localized mainly in the endoplasmic reticulum and mitochondria of liver and other tissues in organisms (Arinc et al., 2000). One of the most widely studied CYP isozymes is CYP1A1, because it is involved in metabolism of large numbers of cytotoxic, carcinogenic, and mutagenic xenobiotics. CYP1A1-associated enzyme activity has been determined by 7-ethoxyresorfin as a substrate. The measurement of 7-ethoxyresorufin O-deethylase (EROD) activity appears to be the most sensitive and the most widely used catalytic probe for determining the induction response of CYP1A in organism (Arinc et al., 2000; Østby and Åse, 2002). EROD is used as a biomarker of exposure to polycyclic aromatic hydrocarbons (PAHs) and structurally related compounds, regarded as aryl hydrocarbon receptor (AhR) ligands (Marohn et al., 2008). However, it has been shown that fish liver microsomal EROD activity maybe inhibited by heavy metals (Sen and Semiz, 2007). Any influence of metals on the capacity of AhR ligands to induce AhR-regulated genes will influence the carcinogenicity and mutagenicity of the AhR ligands (Elbekai and El-Kadi, 2004). Thus, there is a need to understand the mechanisms driving these responses to

∗ Corresponding author. Tel.: +86 532 8289 8583. E-mail address: [email protected] (J. Song). 1382-6689/$ – see front matter © 2010 Elsevier B.V. All rights reserved. doi:10.1016/j.etap.2010.03.008

define the exact role of AhR ligands and metals in carcinogenicity and mutagenicity. The heavy metals discharged to the aquatic environment are of great concern all over the world. Natural and anthropogenic sources such as mine washing or agricultural leaching has resulted in severe heavy metal pollutions in aquatic entrainment, such as lead (Pb), mercury (Hg), and copper (Cu) (Ferreira-Cravo et al., 2009). Heavy metal ions exert their toxicity by multiple mechanisms. Numerous studies showed that exposure to these metals is accompanied by the induction of oxidative stress (Valko et al., 2005). Antioxidant systems have great potential to shed light on the consequences of metal exposures because they represent both coping mechanisms and potential targets. For example, many metals are known to generate oxidative stress either through direct generation of reactive oxygen species (ROS) or by scavenging thiols (glutathione and cysteine) that act as important nonenzymatic antioxidants. On the other hand, enzymatic antioxidants may be susceptible to metal exposure via metal interaction with sulfhydryl or other functional groups, or via the replacement of cofactors essential to enzyme function (Xie et al., 2009). Studies indicated that ROS have been shown to play a role in the decrease of CYP450 in vivo and in vitro, an effect that is modulated by changing its cellular redox potential (Elbekai and El-Kadi, 2005). Metals such as Cd(II), Cu(II), and Pb(II) are known to inhibit the cytochrome P450 enzyme activities (Korashy and ElKadi, 2004) and to reduce the microsomal cytochrome P450 content (Brüschweiler et al., 1996) in mammals. This may be relevant to

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Fig. 1. The location of the stations in Bohai Bay, 2008.

the metabolism of xenobiotics and endogenous substrates in fish (Snyder, 2000), which is of great importance for the environmental toxicity study. Some papers studied the regulation of CYP1A1 by metals. However, the mechanisms responsible for the variation of CYP1A1 activity are still not well known. It should be mentioned that studies focused on modulation effects of heavy metals on CYP1A1 activity were carried out with high exposure level. None of the paper, to our knowledge, has discussed with environmentally relevant concentrations, which is much closer to the actual situation. In our studies, we investigated the pollution status of heavy metals (lead, mercury, and copper) in Chinese Bohai Bay seawater to obtain the concentrations of these metals in natural environment. And then, in lab, bivalves (Chlamys farreri) were exposed to these heavy metals at the level of concentrations obtained from our in situ investigation, in order to find out whether there have been any oxidative stresses caused by these metals at such a low concentration, and whether they would affect the CYP1A1 express. 2. Materials and methods 2.1. Study area The Bohai Sea is a semi-enclosed epicontinental sea of the northeast China and consists of Liaodong Bay, Bohai Bay, Laizhou Bay and the middle sea. Bohai Bay is located in the west of Bohai Sea accounting 20% of the extent of the Sea and with a mean depth of 12.5 m, which is a typical semi-enclosed coastal bay and has limited water exchange with ocean (Meng et al., 2008). It receives a vast amount of fresh water from the rivers around the bay (see Fig. 1), among which Haihe River is the major fresh water source discharging directly to Bohai Bay from the top of the bay’s west. Rapid development of industry has caused serious pollution problems, which have adversely affected the water quality. To describe the distributions of heavy metal in Bohai Bay, seawater samples were collected from 20 stations (between 38◦ 14 04 N and 38◦ 58 00 N), located to account for terrestrial discharge outlets close to shore (see Fig. 1). 2.2. Sample collection and heavy metal determination Water samples from surface seawater of Bohai Bay were manually collected from 20 sampling stations (shown in Fig. 1) during 21–28 April 2008. The samples were taken using 10 L Niskin bottles and filtered through nitric acid pre-cleaned cellulose acetate membrane filters (0.45 ␮m), acidified to a final pH < 2 for dissolved metal analysis, and stored in the dark at 4 ◦ C in 1 L acid-cleaned polyethylene bottles. Lead and copper concentrations were obtained by atomic absorption spectrometry, while mercury concentrations were determined by Cold Vapor Atomic Absorption Spectrometry. All procedures for water sample collection, preservation, pretreatment and analysis were conducted according to the Chinese Specification for Marine

Monitoring (GB17378 2007). Limits of detection were 0.03 ␮g/L, 1.0 × 10−3 ␮g/L and 0.2 ␮g/L for lead, mercury, and copper, respectively. Analytical precision was in good agreement, generally better than ±5.0% relative standard deviation. The recovery values of metal analysis were between 90% and 110%. 2.3. Animals and treatments The bivalves (C. farreri) were collected during November from a C. farreri farm in a clean area of Shandong provinces. The bivalves were transferred to a flow through cultivate system, with fresh seawater (11 ◦ C) for an acclimation period for 3 days. Bivalves (selected for size 5–6 cm length) were randomly divided into four experimental groups. One was the control group and the other groups were exposed to nominal lead, mercury and copper concentrations of 2 ␮g/L, 5 × 10−2 ␮g/L, and 3 ␮g/L, respectively (PbCl2 , HgCl2 , and CuCl2 , all from Sinopharm, China). Each group has 9 individuals, and both control and exposure groups were under the identical conditions, with 8 L of well-aerated pre-treated clean sea water (the seawater was treated through precipitation process in sedimentation tank, and then through granular activated carbon filtration), the temperature was maintained at 10 ± 2 ◦ C, during experiment. The aquarium sea water was refreshed for 2 L every day due to the rejection of bivalves and the metal depletions, and to keep the metals at relative constant levels. After 96 h exposure, the bivalves were killed for analysis. During experiment, no dead bivalves were found. For the enzymatic assays, digestive glands were homogenized (1:10, w/v), in ice-cold buffer with pH adjusted to 7.60 (50 mM phosphors buffer, 1 mM DDT, 1 mM EDTA and 150 mM KCl). Homogenates were centrifuged at 9000 × g for 20 min (4 ◦ C) to remove cell debris and then postmitochondrial supernatant (S9) from both control and exposure groups were collected and stored at −80 ◦ C until employed later to determine total protein content, enzymatic activities (including superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx), and glutathione S-transferases (GST)) and lipid peroxidation (MDA). The postmitochondrial supernatant was spun down for 1 h at 105,000 × g (4 ◦ C) in an L80K ultracentrifuge (Beckman Instruments, USA). The microsomal pellet was resuspended in 2 ml of buffer (pH 7.60) and the suspension was further diluted in the same buffer (1:10 or 1:20) and stored at −80 ◦ C until used for EROD determination. All the determination assays were performed at least triplicate. 2.4. Biochemical assays Digestive gland EROD activity was determined according to a modified method by Kennedy and Jones (1994). Phosphate buffer (31 ␮L, 50 mM, pH 7.6) was added to each well of 96-well plate. Each well then received microsomal suspension (30 ␮L) and ethoxyresorufin (50 ␮L with final concentration 10 ␮m). The reaction solution was incubated at 25 ◦ C for 10 min, and the reaction was started with the addition of NADPH (17 ␮L with final concentration 1.87 mM), to each reaction well. Blank wells received phosphate buffer (17 ␮L, 50 mM, pH 8.0). The plate was then immediately placed in a multi-well plate reader (BioTek Instruments, Synergy 2 Multi-Mode Microplate Reader, USA) and the resorufin fluorescence was monitored for 30 min by repeated measurements at 544 nm (ex) and 590 nm (em). After the last measurement the plate was removed from the reader, 72 ␮L of fluorescamine in acetonitrile (250 ␮g/ml) was added to each well to stop the reactions and the plate was incubated with continuous shaking and protected from light at 25 ◦ C, 15 min, the fluorescence

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of proteins was read at an excitation wavelength of 390 nm and an emission wavelength of 460 nm. A standard curve of resorufin was used to read the concentration of the EROD production, and bovine serum albumin (BSA) standard curve was used for that of total protein, to see the detail of the standard curve please refer to the protocol described in Kennedy and Jones (1994). CAT activity of tissues was determined according to the method of Aebi (1984). The enzymatic decomposition of H2 O2 was followed directly by the decrease in absorbance at 240 nm. The difference in absorbance per unit time (30 s in our assay) was used as a measure of CAT activity. The enzyme activities were given in U/mg protein. SOD, GPx, GST and lipid peroxidation (MDA) were assayed by using kits from NanJing JianCheng (NanJing JianCheng Bio Inst., China). All enzymes were normalized to total protein, which was determined estimated using Coomassie Brilliant Blue (G-250) by the method of Bradford (1976) using a BSA standard, except the protein measured in EROD assay, which the enzyme source is different and the protein have been measured by fluorescence plate reader as mention above. 2.5. Data statistical analysis All data are expressed as the means ± SD. To identify significant differences obtained by one-way ANOVA, pair wise comparisons between experimental groups and the control group were performed by Dunnet’s post hoc method, p < 0.05 was considered to be statistically significant. Statistical analyses were performed using SPSS 13.0 software (SPSS, Chicago).

3. Results Distribution patterns of lead, mercury, and copper concentrations have been shown in Fig. 2. The concentrations of these metals in surface seawater of Bohai Bay did not vary a lot with stations, ranging from 1.25 ␮g/L–2.02 ␮g/L, 3.48–6.83 × 10−2 ␮g/L, and 2.22–3.12 ␮g/L, with the mean values of 1.63 ␮g/L, 4.85 × 10−2 ␮g/L, and 2.68 ␮g/L, for lead, mercury and copper, respectively. The concentrations of heavy metals were highest near the Haihe River (A1 or A2 stations), but showed no obvious geographical differences. The concentrations of mercury were comparable with that of the first class limit of seawater quality standard in China (GB3097, 1997), while lead and copper were some extent higher and lower than that. When bivalves exposed to the environmentally relevant concentrations of heavy metals, the antioxidant enzymes showed different responses. The results of antioxidant enzyme activities (SOD, CAT, and GPx) were shown in Fig. 3. The reduced SOD activity under lead exposure could be detected by 8.8% reductions, the other two metals had statically non-significant effect on the activity of SOD (p > 0.05). For the change of CAT, the activities were both reduced in lead- and mercury-exposed groups by 28.7% and 9.1%, respectively; Exposure to copper did not bring about an appreciable change in CAT activity. Depletion in GPx activity were observed in all of the exposed groups, exhibiting only 37.6%, 66.7% and 91.1% of that in control group for lead, mercury and copper, respectively. It is noted that the GPx activity in copper-exposed group was marginally but not greatly lower than that of control. In respect to GST, all of the three metals seemed to show a positive effect on it. There was no statistic difference between all exposed groups and control (p > 0.05). MDA content was observed significantly increased in the lead-exposed group. However, results showed no clearly lipid peroxidation after exposing to mercury or copper. For EROD activity measured in bivalves exposed to lead, a 36.9% reduction was observed as compared to that of control group (Fig. 3). For the other metals no significant variation of this enzyme was detected as compared to that of control group. 4. Discussion 4.1. Distribution of lead, mercury and copper in seawater of Bohai Bay Dissolved lead, mercury and copper concentrations of 20 samples from surface seawater of Bohai Bay were detected. The

Fig. 2. Distribution of Cu, Hg, and Pb in surface sea water of Bohai Bay. The regulatory guideline is refer to the first class limit of seawater quality standard in China (GB3097, 1997).

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Fig. 3. Digestive gland enzymatic activity and MDA content in bivalve C. farreri exposed to Pb, Hg and Cu at environmentally relevant concentrations (2 ␮g/L, 5 × 10−2 ␮g/L, and 3 ␮g/L, respectively). Data expressed as average ± SD (n = 9), ‘*’ indicates that the value is significantly different (p < 0.05) from that of the control.

distribution maps of three metals were shown in Fig. 2. The distribution patterns of the three metals were similar to each other. The concentrations did not vary a lot with the locations. Although, the highest concentrations were detected near the Haihe River (A1 or A2 stations) for three metals, there were no obvious variation patterns which could be drawn. The variation pattern of cadmium described in our previous study, the concentrations of cadmium in the inner part of Bohai Bay were higher than those in the outer part. Taking into consideration the relationship between cadmium concentrations and the downstream distances of Haihe River, it shows that the concentrations decreased with increasing distance from the shore. However, unlike the co-change with distance found in cadmium, none of lead, mercury, nor copper showed this variation pattern. It is because that the lead, mercury can be transported to the off shore region via atmospheric deposition (Lacerda and Ribeiro, 2004), while, cadmium is mainly from riverin input. Copper is a trace element essential to life (Ferreira-Cravo et al., 2009), it may be taken by the biomass. Therefore, the distribution pattern would be influenced not only by the river input but also by the biomass distribution.

Bohai Bay had been considered to be contaminated severely in the last 20 years. According to published data, although the Bohai Sea makes up only 1.6% of the total Chinese sea areas, it receives about 36% of the wastewater and 47% of the solid pollutants in China (Peng et al., 2009). In 2001, China launched its first regional ocean governance program, the Clean Bohai Sea Program, for the purpose of cleaning up its most polluted sea area. However, the effectiveness of the program has still been contested (Peng et al., 2009). In our study, the mean concentrations of lead, mercury and copper were 1.63 ␮g/L, 4.85 × 10−2 ␮g/L, and 2.68 ␮g/L, respectively. Compared with another investigation carried out in 2003 (Meng et al., 2008), mercury and copper did not change a lot in the past 5 years, from 0.04 ␮g/L to 4.85 × 10−2 ␮g/L and from 2.54 ␮g/L to 2.68 ␮g/L, respectively. However taking a look at the variation of lead, the situation have turned better, the concentration used to be 7.18 ␮g/L (Meng et al., 2008) but has dropped to 1.63 ␮g/L in our study. It was noteworthy that only copper in all of the stations we investigated in seawater of Bohai Bay were lower than the first class limit of seawater quality standard in China (lead, 1 ␮g/L, mercury, 0.05 ␮g/L, and copper, 5 ␮g/L) (GB3097, 1997). However, the

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concentrations of lead in all of the samples and that of mercury in 7 out of 20 samples exceeded the first class limit. 4.2. Biomarker investigation Previous reports showing that the heavy metals affected the induction of CYP1A1 prompted us to investigate the mechanisms of this interaction, to enhance our understanding of how these metals could alter the activity of CYP1A1. Concretely, we set experiments to examine the role of metal-induced oxidative stress on the expression of CYP1A1. We initially checked the potential of the metals, at the level of environmentally relevant concentrations, to induce oxidative stress in bivalves (C. farreri) using oxidative system biomarkers: the activities of antioxidant enzymes (SOD, CAT, GPx, and GST), and induction of lipid peroxidation (MDA). Next, we determined the effect of the oxidative responses on the CYP1A1 activity by these environmentally relevant concentration levels heavy metals. 4.3. Antioxidant enzyme system Copper did not obviously affect the antioxidant system in our study. The changes in SOD, CAT, and GST were statically non-significant (p > 0.05). Even the inhibition on GPx considered statistically significant, has inhibitory rate of only 8.83%. Copper was reported to be a toxic metal (Ferreira-Cravo et al., 2009; Geracitano et al., 2002). It is generally accepted that copper toxicity is a consequence of the generation of ROS by copper ions via Fenton or Haber–Weiss reactions (Letelier et al., 2005). Although, it have been suggested that copper ion displays high affinity for thiol and amino groups occurring in proteins (Letelier et al., 2005). Thus, specialized proteins containing clusters of these groups transport and store copper ion, hampering its potential toxicity. This mechanism may be overwhelmed under copper overloading conditions (Alexandrova et al., 2007), and turn out to induce toxicity via binding to functional group of protein. Therefore, we find such numerous studies indicating the toxicity of copper, since most of them were carried out in laboratories, in which organisms may be exposed to high level of copper to obtain a toxic model, some of which may be up to 0.5 mg/L (Geracitano et al., 2002). In our study, we chose the average of concentrations of copper in Bohai Bay seawater as the exposure level, which was even lower than that of the first class limit of seawater quality standard in China, identified as no pollution. Therefore, it is acceptable that copper exposure in this study showed no obvious toxic effect on bivalves. Inorganic mercury is one of the dangerous environmental pollutants recognized as a highly toxic compound (Bando et al., 2005). While in our study, mercury did not show clearly toxic effect. As the same situation as that of copper-exposed group, which was mainly because of the concentration used in our experiment was only 5 × 10−2 ␮g/L, quite lower than those of other high level exposure. Among all enzymes investigated in our study, only GPx was significant inhibited in mercury-exposed group, by 33.2%. The inhibition of GPx caused by mercury may be due to its high affinity for thiol. Inorganic mercury binds to various molecular weights of thiol-containing proteins (glutathione, cysteine, and albumin) with high binding ability. Mercury competes for thiol groups on the GSH molecule which forms GS–Hg–SG from GSH (Becker and Soliman, 2009), and reported to have resulted in a highly significant (p < 0.01) decrease in adult fish hepatic GSH contents (Allen et al., 1988). Depletion of intracellular thiol through binding with mercury can alter the nature and activity of proteins within cells, possibly contributing to oxidative stress (Becker and Soliman, 2009). Thus, reduced supply of GSH could be a cause for the reduced GPx activity (Jadhav et al., 2007).

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Lead was much more effective than copper and mercury on activities of antioxidant enzymes. It has been reported that the interference of lead resulted in the generation of highly ROS, including O2 •− , H2 O2 , and • OH (Bokara et al., 2008). SOD, CAT, and GPx has been shown to play an important role in protection against oxidation. However, the excess availability of free radicals may decrease the levels of antioxidants. SOD are a group of metalloenzymes that catalyze the conversion of O2 •− to yield H2 O2 , which itself is an important ROS as well (Li et al., 2009). The results of researches on the lead influence on the SOD activity are divergent. When pearl oysters exposed to 0.5 ␮M lead, the mantle SOD activity of lead treatment was higher than that of control throughout the period of exposure (Jing et al., 2007; Adonaylo and Oteiza, 1999) did not notice any changes in the SOD activity in rats. Whereas, our results were compatible with the researches of many authors who found decreased SOD activity after lead exposure. Decreased activities of antioxidant enzymes were reported in the liver of castrated boars (Yu et al., 2008), in kidney of rats (Patra et al., 2000) and in all regions of all the treated Wistar dams (Babu et al., 2007). This may be caused by interaction between lead and copper, a metal necessary for the proper functioning of the SOD enzyme. Lead exposure has an inhibitory effect on SOD, thus may cause alterations in mitochondria facilitating the release of O2 •− , which has been shown to inhibit CAT activity (Hansen et al., 2007). As showed in Fig. 3, lead exposure has caused 28.7% reductions in activity of CAT. It has been suggested that decrease of CAT activities were attributed to the inactivation of enzyme or change in assembly of enzyme subunits, the decline in CAT activities under lead exposure suggested a possible delay in removal of H2 O2 and toxic peroxides and favors H2 O2 accumulation mediated by ROS (Wang et al., 2008). Exposed to lead has significantly inhibited the activity of GPx when compared with control. As the same as mercury, lead is know to inhibit several antioxidant enzymes having functional sulfhydryl groups (Patrick, 2006), which indirectly reduced the GPx activity. The decreased activities of antioxidant enzymes observed in the present study may be due to over production of reactive oxygen metabolites (ROMs) especially O2 •− by lead. Another important reason for the decreased activity of the antioxidant enzymes could be attributed to their nature of synergistic functioning, which may at least partly explain the mechanism of their reduced activity. A decrease in SOD activity increases the level of O2 •− , which is known to inactivate CAT activity (Kono and Fridovich, 1982); when CAT or GPx fails to eliminate H2 O2 from the cell, the accumulated H2 O2 causes inactivation of SOD (Jadhav et al., 2007). 4.4. Lipid peroxidation Change in activities of SOD, CAT, GPx, and SOD were associated with significant change in MDA. As Fig. 3 shows, the MDA level of lead-exposed group was higher than the control values, and increased by 23.59%. These findings agreed quite well with previous investigations showing that in vivo, lead-induced lipid peroxidation results in the formation of aldehydic by-products such as MDA (Caylak et al., 2008). We suggested that the potential mechanism for lead intoxication in digestive gland as follows: lead caused decrease in activities of antioxidant enzymes, leading to the declined ability of scavenging free radicals with excessive production of lipid peroxides, resulting peroxidative damage to membrane lipids. 4.5. CYP1A1-dependent monooxygenase activities Metals are known to inhibit CYP1A1-dependent monooxygenase activities (Pereira et al., 2009). According to our results, we found that lead was generally the most potent in the inhibition of tested antioxidant enzyme activities (SOD, CAT and GPx), in high level of ROS and consequent lipid peroxidation increase. Accompa-

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nying that, lead-exposed group also induced a marked inhibition on EROD. We speculated that there are some relationships between oxidative stress and CYP1A1 activities. It have been documented that CYP1A genes can be responsible to oxidative stress and accumulation of mRNA is prevented by ROS at the transcriptional level (Benedetti et al., 2007); ROS and in particular H2 O2 , are also known to down-regulate CYP1A expression acting in the signaling pathway through the stress response transcription factor, nuclear factor I (NFI) and nuclear factor kB (NF-kB) (Barouki and Morel, 2001). In our research, CAT and GPx, the antioxidant enzymes that act on H2 O2 , have been inhibited by lead. It may be one of the reason accounting for the decline of EROD activity, since it has been documented that the protein content and the catalytic activities of CYP enzymes can be reduced when ROS act as second messengers leading to activation of kinases with phosphorylation of CYP450 proteins or when they interact with the iron of heme groups (Benedetti et al., 2007). Heavy metals were verified to reduce considerable structural and functional change of protein and alter activity of enzyme by binding to their functional group (sulfhydryl, carboxyl, imidazol, etc.) or by displacing the metal associated with enzyme (Rissode Faverney et al., 2000). Studies in 1970s have established that lead was an inhibitor of a number of cytochrome P450 related oxidations (Alvares et al., 1972). The reasons for this could be sought in the combined effect of this metal on protein. Previous researches showed that lead can inhibit ␦-aminolevulinc dehydratase, which is an important enzyme in heme synthesis pathway, via binding of the lead to thiol groups of allosteric sites provoking allosteric transitions to inactive form of enzyme, leading to a decrease in the heme synthesis (Brüschweiler et al., 1996; Korashy and El-Kadi, 2004). The hematological system is the major target of low level lead exposure, the same decrease phenomena have been found in other researches (Caylak et al., 2008). Since heme is the prosthetic group of CYP450 (Korashy and El-Kadi, 2004), reducing of the bioavailability of this prosthetic group has been indicated as an important pathway for decreased levels of CYP1A1 proteins accompanied by an inhibition of CYP1A1 activity. Another potential reason may be involved in the inhibition of CYP1A1 activity by lead via altering the structure of the phospholipids. A direct interaction of lead with biological membranes, inducing lipid peroxidation was recorded in previous studies, which may alter the structure of phospholipids (Moreira et al., 2001), thereby affecting lipid organization and electron transport among microsomal CYP components (Sen and Semiz, 2007), indirectly, leading to the inhibiton of CYP1A1 activity.

5. Conclusions In summary, the present study suggested that the concentration levels of lead, mercury, and copper in the surface seawater were around the first class limit of seawater quality standard in China. Compared with other investigations carried out in 5 years before, the water quality in respect to lead had turned better, while mercury and copper did not change a lot in the past 5 years. Although, the lead showed an improving trend, it was the only metal in our investigation that has exceeded the first class limit of sea quality standard in China. Based on our in vivo experiments, it is recommended that exposure to environmentally relevant concentration of lead induced several biochemical responses in antioxidant enzymes, resulting in lipid peroxidation. At the mean time, CYP1A1 enzyme activity has been inhibited, which may have some relationships with the oxidative stress. Multi-biomarker responses in bivalves when exposed to lead at

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