AHH activity, tissue dose and DNA damage in different tissues of the scallop Chlamys farreri exposed to benzo[a]pyrene

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Environmental Pollution 153 (2008) 192e198 www.elsevier.com/locate/envpol

AHH activity, tissue dose and DNA damage in different tissues of the scallop Chlamys farreri exposed to benzo[a]pyrene Luqing Pan*, Jingjing Miao, Jing Wang, Jing Liu The Key Laboratory of Mariculture, Ministry of Education, Ocean University of China, No. 5, Yushan Road, Qingdao, China Received 27 October 2006; received in revised form 25 June 2007; accepted 13 July 2007

The combination of AHH activity, B[a]P accumulation and DNA damage in the scallop Chlamys farreri can be a reasonable biomarker battery for pollution monitoring. Abstract A collaborative study was performed on scallops (Chlamys farreri) exposed to 0.5, 3 and 10 mg/L benzo[a]pyrene (B[a]P) for 20 days. The levels of aryl hydrocarbon hydroxylase (AHH) activity, B[a]P accumulation and DNA strand break were assayed in the gill and digestive gland. Results showed that AHH activity and B[a]P accumulation were significantly related to B[a]P dose. AHH activity was induced and then became stable gradually. B[a]P accumulation increased first and showed an incoming plateau. DNA strand break levels in the 0.5 and 3 mg/L B[a]P groups remained high and significantly different from control values until day 6, followed by a reduction in the gill and no recovery in the digestive gland. The 10 mg/L B[a]P groups remained significantly lower than control until the end. These results suggested that the application of comprehensive indices may give information on the actual exposure of organisms to pollutants, and also information on toxic effects. Ó 2007 Elsevier Ltd. All rights reserved. Keywords: Benzo[a]pyrene; Chlamys farreri; Aryl hydrocarbon hydroxylase; Bioaccumulation; DNA strand breaks

1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are common coastal environmental organic pollutants derived from petroleum, ocean shipping and industry sewage. Owing to their toxicity (highly mutagenic and carcinogenic features) and widespread distributions in the global environment, PAHs have attracted the attention of environmental chemists, toxicologists and regulatory agencies (Menzie et al., 1992). In China, coastal PAHs pollution monitoring has been traditionally based on chemical analysis of contaminant levels in sediment (Fan and Cote, 1990; Mai et al., 2001). More recently, bivalves have been proved to be suitable bioindicators for monitoring trace toxic contaminant levels of coastal waters; i.e., Mussel Watch in the United States and RNO in France (Goldberg,

* Corresponding author. Tel.: þ86 532 8203 2963; fax: þ86 532 8289 4024. E-mail address: [email protected] (L. Pan). 0269-7491/$ - see front matter Ó 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.envpol.2007.07.022

1978; RNO, 1988), because they take up and concentrate contaminants to levels above those in the surrounding seawater, and the examination of the tissue dose gives information on the bioavailable fractions which may cause deleterious effects. Meanwhile bivalves also exhibit a certain ability to respond to pollutants at a molecular level in the first place. So presently, attention has focused on a contaminant monitoring strategy that includes not only the analysis of contaminant levels in biota but also the detection of biological responses of bivalves. The strategy can give information on the actual exposure of organisms to pollutants, and also information on toxic effects (Huggett et al., 1992; Walker et al., 1997; Anderson et al., 1999; Quinn et al., 2005). It has been established that once the PAHs enters the organism’s system, it can experience a series of oxidative processes mediated by the mixed-function oxygenase (MFO) system (phase I). The phase I by-products may undergo further metabolism by conjugative enzymes (phase II) such as glutathione S-transferase (GST), which will make them more readily

L. Pan et al. / Environmental Pollution 153 (2008) 192e198

excretable. However, these reactions might bring about the synthesis of more reactive molecules that can interact with the genetic material and cause DNA damage. They may also increase the presence of reactive oxygen species (ROS) that can be responsible for lipid peroxidation processes and result in DNA strand break eventually (Stansbury et al., 1994; Cavalieri and Rogan, 1995). Although bivalves appear to be suitable bioindicators, their biochemical pathways are distanced from vertebrates’ and we still need information on mechanisms underlying xenobiotic uptake and metabolism (Kurelec et al., 1996; Chaty et al., 2004). As a most important PAHs metabolism enzyme system, CYP450 dependent response has become the most commonly used indicator of PAHs exposure in fish and the induction can be assessed by measuring EROD and aryl hydrocarbon hydroxylase (AHH) activities (Collier and Varanasi, 1991; Van Veld et al., 1990). Moreover, there are many published results that have indicated that CYP450 monoxygenases or MFO are present in bivalves, although to a limited extent if compared with fish and mammals (Willett et al., 1999; McElroy et al., 2000; Pan et al., 2005). Many studies have reported that AHH activities in the digestive gland and gill, respectively, were elevated in mussels collected at sites more highly contaminated with PAHs (Willett et al., 1999; McDonald, 1990; Michel et al., 1994; Garrigues et al., 1990). In contrast, Willett et al. (1999) reported that no detectable induction of AHH was observed in mussels injected with B[a]P (5 mg/kg) or TCDD (20 mg/kg) for 48 h. Therefore, further research is necessary to determine the time-course and dose-dependent relationship of AHH enzymes in bivalves exposed to PAHs. On the other hand, because of the comparatively low-metabolism capacity as well as the filter-feeding trait of bivalves, a great deal of PAHs cannot be metabolized and accumulate in the their tissue. Many field and laboratory studies have reported that PAHs dose in mussels can sensitively reflect the environmental pollution level, and a doseeresponse pattern of B[a]P accumulation in mussels has been reported (Canova et al., 1998; Skarphe´ðinsdo´ttir et al., 2003), but little work has been done on the time-course of B[a]P accumulation in tissues. Meanwhile, as the PAHs are metabolized, reactive molecules are produced and interact with the genetic material and cause DNA damage as a toxic terminal of the induction of a cascade of cellular events. DNA integrity reduction represents the genotoxicity which may have a long-term effect on the sustainability of a particular population. DNA strand breakage in bivalves using alkaline unwinding has been approved as an effective biomarker to assess the genotoxicity of pollutants (Everaarts and Sarkar, 1996; Ching et al., 2001; Siu et al., 2003). The present study of DNA strand break is an extension of the previous work in order to have a comprehensive understanding of the DNA toxicity. It is well known that bivalves are a very diverse group that have adapted to a large number of lifestyles, mainly including infaunal and epifaunal. Considering that contaminant levels of PAHs are different in sediment and in seawater, it is thereby necessary to adopt bivalves of different lifestyles to monitor their habitat pollution. The scallop Chlamys farreri is a main

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commercial mollusk widely cultured by rope-growing method in the shallow seas of China. In the present study, we used benzo[a]pyrene (B[a]P), a powerful carcinogen among the 16 PAHs considered as primary contaminants (Menzie et al., 1992), as a model substance and undertook a collaborative study to evaluate in parallel the tissue B[a]P accumulation, AHH activity, as well as DNA strand break in the scallop C. farreri. Our specific objectives in this study were to compare systemically the temporal and doseeresponse relationship between B[a]P exposure and toxicological parameters including B[a]P accumulation, AHH activity and DNA damage in different tissues (gill and digestive gland), as well as to examine which tissue is more susceptible to B[a]P exposure. This study may provide scientific data for development of the contaminant monitoring strategy that has been mentioned initially, and the bioaccumulation results may make sense for food security as C. farreri is a commercial species. 2. Materials and methods 2.1. Chemicals Benzo[a]pyrene (Fluka, USA) was dissolved in acetone. All chemicals for sample preparation and HPLC detection were of chromatogram grade. Biochemicals for AHH activity analysis and alkaline unwinding were of analytical grade.

2.2. Experimental organisms Scallops, C. farreri, aged 2 years and with shell length of 6.0  0.5 cm, used for this experiment were rope-growing scallops, collected from the Pacific Corner (Yellow Sea, Qingdao, China). Scallops in glass tanks (1 L water per mussel) were acclimated to photophobic underground laboratory conditions for 7 days in filtered seawater pumped from the Bay of Pacific Corner (Yellow Sea, Qingdao, China). The seawater was continuously aerated, and salinity, temperature and pH were maintained at 30&, 10 (2)  C and 8.1, respectively. One third of the water was renewed every day. The scallops were fed with dried powder of Spirulina platensis (30 mg for each individual) daily.

2.3. Experimental design We have tested the PAHs concentration in seawater from the Bay of Pacific Corner by HPLC before the experiment, and the concentration of the B[a]P was 0.155 ng/L. The water solubility of B[a]P was 16 nM (Schirmer et al., 1998). Therefore, to ensure that the nominal B[a]P concentration be stable, 0.5, 3 and 10 mg/L were chosen as waterborne B[a]P concentrations. B[a]P was first dissolved in acetone, and then added to seawater to achieve a final acetone concentration of 0.0025%. So control groups include a seawater control and an acetone control. There were three tanks used as replicates per group. Experimental conditions (salinity, pH, temperature, scallop density, feeding) were the same as those used for acclimation, and one third of the exposure media was renewed daily. The exposure experiment lasted for 20 days. During this period, we determined the B[a]P concentration of exposure groups every day before renewing the water during the experiment. The analyzed B[a]P concentrations were 0.48  0.08, 3.12  0.10 and 9.06  0.16 mg/L. Gills and digestive glands were excised and frozen immediately at 80  C for subsequent examination.

2.4. Aryl hydrocarbon hydroxylase activity assays Scallops were sampled at days 0, 0.5, 1, 3, 6, 10, 15 and 20 for the AHH activity assay; two to three scallops were used per replicate. The gills and digestive glands of scallops, respectively, were homogenized in a buffer containing

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20 mM TriseHCl, 1.5 mM EDTA, 1.0 mM dithiothreitol, and 10% glycerol (v:v) (pH 7.6). Samples were centrifuged for 25 min at 600 g, 20 min at 13 000 g and 70 min at 100 000 g. The resulting microsomal pellet was resuspended in a 50 mM TriseHCl buffer containing 1 mM EDTA in 20% glycerol. Microsomal protein was quantified with a commercial kit (Bio-Rad) based on Coomassie blue, using bovine serum albumin standards (Bradford, 1976). Aryl hydrocarbon hydroxylase activity (AHH) was determined as described by Willett et al. (1999) with modifications. The 1.0 mL incubation mixture consisted of 50 mM TriseHCl, pH 7.6, 0.1 mM NADPH, and 500 mg of microsomal protein. Samples were preincubated at 30  C and the reaction was initiated by the addition of 60 mM B[a]P. Samples were incubated for 30 min and stopped with the addition of 1 mL cold acetone followed by 3.25 mL hexane. Samples were vortexed, 2 mL of the organic layer was drawn and extracted with 5 mL aqueous NaOH, and fluorescence was determined with a spectrofluorometer (Molecular Spectroscopy LS 55 from Instruments, P.E., MA, USA) at 396/522 nm (excitation/emission). The spectrofluorometer was calibrated using an authentic 3-OH B[a]P standard.

2.5. Tissue B[a]P analysis and fat analysis Scallops were sampled at days 0, 3, 10 and 20 for tissue B[a]P analysis. Gills and digestive glands were analyzed following cyclohexane extraction and HPLC/fluorimetric detection. About 20e30 g of gills and digestive glands, respectively, from 20 scallops were homogenized. After measuring the wet weight, samples were saponified with 80% KOH and 20 mL alcohol at 60  C. Three extractions with 30 mL of cyclohexane were performed, and the combined extracts were washed with KOH 10% and saturated NaCl. Then, the volume of the extract was reduced to 1 mL and the solvent exchanged with methanol under a gentle nitrogen stream. After dilution to 10 mL, samples were injected through a syringe-filter into the chromatograph and isocratic elution (water:methanol 20:80 v/v) at 1 mL/min was applied. The HPLC system was equipped with an integrator (Agilent 1100, USA), a ZORBAX ExtendC18 column, 5 mm, 250  4.6 mm (Agilent, USA) and an FID detector (Agilent 1100, USA) set at 280/425 nm. A guard column (Hypersil C18 7992080-504, Agilent, USA) was employed to avoid the introduction of strongly retained compounds into the analytical column. The use of a multi-port valve between the guard and the analytical column allowed on-line back washing of the former during the analysis through a supplementary pump, which provided a flow of 80% methanol. B[a]P was quantified by single-point calibration while the assessment of B[a]P recovery was based on the internal standard addition method. B[a]P concentrations were then calculated on wet weight basis. Before the exposure experiment, about 20e30 g of gills and digestive glands, respectively, excised from 20 to 25 scallops were homogenized. After measuring the wet weight, lipid content was determined by extraction with a chloroformemethanol mixture according to Bligh and Dyer (1959). Lipid contents were then calculated on wet weight basis.

2.6. Alkaline unwinding assay Scallops were sampled at days 0, 0.5, 1, 3, 6, 10, 15 and 20 for alkaline unwinding assay and two to three scallops were used per replicate. DNA was extracted from gills and digestive glands by mashing the tissues with plastic sticks in 3 mL TE buffer (50 mM Tris, 100 mM EDTA, pH 8.0). 0.45 mL lysate was transferred into 1.5 mL Eppendorf tubes. Then 0.05 mL of 10% SDS and 3 mL of 20 g/L proteinase K were added and the mixture was incubated for 5 h at 55  C. An equal volume of buffered phenol/chloroform/isoamyl alcohol (PCI) (25:24:1, v/v/v, pH 8.0) was then added to the sample. The sample was gently mixed for 15 min and then centrifuged for 10 min at 13 000 rpm at 4  C. The aqueous layer was transferred to a new tube and digested with 0.8 mL RNase (10 mg/mL) for 30 min at 37  C and the digesta was extracted twice by PCI. DNA was precipitated from the resulting aqueous layer by adding 2 volumes of absolute ethanol and 1/10 volume of 3 M sodium acetate, pH 5.2. The samples were allowed to settle for 2 h at 20  C, and then centrifuged for 15 min at 13 000 rpm. The resulting pellet was rinsed with 1 mL 70% ethanol and air-dried, then dissolved in 1 mL of TE buffer (10 mM Tris, 1 mM EDTA, pH 8.0). The amount of DNA was quantified using a UV/Visible spectrophotometer (Ultro spec 2100 pro, Amersham Biosciences, Sweden).

The alkaline unwinding assay used in the study was adapted from Ching et al. (2001). In the assay, the rate of transition of double-stranded DNA (dsDNA) to single-stranded DNA (ssDNA) under pre-defined alkaline denaturing conditions was proportional to the number of breaks in the phosphodiester backbone, and thus was used as a measure of DNA integrity (Daniel et al., 1985). The amounts of various types of DNA were quantified by measuring the varying degrees of fluorescence resulting from a DNA-binding dye, bisbenzamide. After reaction with the dye, the fluorescence of dsDNA is double that of the ssDNA (Cesarone et al., 1979). The DNA sample was diluted and divided into three equal portions for fluorescence determination of dsDNA, ssDNA and alkaline unwound DNA (auDNA). The fluorescence of the initial DNA or dsDNA was determined by placing 100 mL of the DNA sample, 100 mL of 25 mM NaCl and 2 mL of 0.5% SDS in a prechilled test tube, followed by the addition of 3 mL of 0.2 M potassium phosphate (pH 6.9), and 3 mL of bisbenzamide (1 mg/mL). The contents were mixed and allowed to react in darkness for 15 min. The fluorescence of the sample was measured using a spectrofluorimeter (Molecular Spectroscopy LS 55 from Instruments, P.E., MA, USA) with an excitation wavelength of 360 nm and an emission wavelength of 450 nm. The fluorescence of ssDNA was determined as above but using a DNA sample that had already been boiled at 80  C for 30 min to completely unwind the DNA. auDNA samples were made by subjecting the initial DNA samples to alkaline treatment, i.e., 50 mL of 50 mM NaOH was rapidly mixed with 100 mL of DNA sample in a prechilled test tube and incubated on ice in darkness for 30 min, followed by a rapid addition and mixing of 50 mL of 50 mM HCl and 2 mL of 0.5% SDS; then the mixture was forcefully passed through a 21G needle several times. The fluorescence of the alkaline unwound DNA sample was measured as described above. The ratio of double-stranded DNA to total DNA (F value) was determined as follows: F value ¼ ðauDNA  ssDNAÞ=ðdsDNA  ssDNAÞ

2.7. Statistical analysis All values are expressed as mean  SD of three replicates in one representative experiment. All experiments were performed three times to confirm the results. The analyses were carried out using SPSS software (Version 13.0). Data were compared with one-way or nested ANOVA, and NewmaneKeuls posthoc test when appropriate. The a level used was 0.05. All data were checked for normality and homogeneity of the variances, and transformed when necessary.

3. Results 3.1. AHH activity The AHH responses of the gill and digestive gland are shown in Fig. 1. The results indicated no significant difference (P > 0.05) of the AHH activities between seawater control groups and acetone control groups during the experimental period. Meanwhile, AHH activities in digestive gland are consistently higher than those in gills. The AHH activities in gill of the 0.5, 3 and 10 mg/L treatment groups increased and became significantly different from control values at the 12th hour (P < 0.05). The AHH activities remained high and significantly different from control values until the end of the experiment, and the AHH activities of the 0.5, 3 and 10 mg/L treatment groups became stable at day 3, day 6 and day 15, respectively (P > 0.05). The AHH activities in digestive gland of the 0.5, 3 and 10 mg/L treatment groups increased and became significantly different from control values at day 3, day 1 and the 12th hour, respectively (P < 0.05). AHH activities of the 0.5 mg/L

AHH activity (pmol mg-1 min-1)

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Gills 12 10 8 6 4 2 0

0

5

10

15

20

3 µg/L

10 µg/L

Time (days)

AHH activity (pmol mg-1 min-1)

Control

Acetone

0.5 µg/L

Digestive glands 12 10 8 6 4 2 0

0

5

10

15

20

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gills and digestive gland of experimental scallops are reported in Table 1. Scallops of control groups showed detectable B[a]P concentrations because the area where we sampled the scallops had a B[a]P concentration of 0.155 ng/L, and the experimental seawater was pumped from the same area. In contrast with the control group, B[a]P dose of gills and digestive glands in scallops exposed to 0.5, 3 and 10 mg/L B[a]P rose parallel with the exposure dose. In gill samples of scallops from the 0.5, 3 and 10 mg/L treatment groups, B[a]P accumulation was constant from day 10, for there were no marked differences among the internal B[a]P doses of days 10 and 20 (P > 0.05). In digestive gland samples of scallops from the 0.5 mg/L treatment group, B[a]P uptake was constant from day 3, for there were no marked differences among the internal B[a]P doses of days 3, 10 and 20 (P > 0.05), while in digestive gland samples from the 3 and 10 mg/L treatment groups B[a]P was constant from day 10. Tissue differences in B[a]P accumulation were observed. B[a]P concentrations in the digestive gland were consistently higher than those in the gill. In order to study the relationship of tissue B[a]P concentration and tissue lipid content, we analyzed the lipid content of the gill and digestive gland as well. Lipid content of the gill was 1.297  0.041%, and lipid content of the digestive gland was 4.940  0.084%.

Time (days)

3.3. DNA strand breaks Control

Acetone

0.5 µg/L

3 µg/L

10 µg/L

Fig. 1. AHH activity of gill and digestive gland of scallops exposed to B[a]P (0.5, 3.0 and 10.0 mg/L) for 30 days compared with unexposed group (n ¼ 3). Mean  SD of three to five determinations in triplicate, AHH activity was induced significantly in treatment groups (P < 0.05).

treatment group at days 6 and 10 had no significant difference (P < 0.05), and then decreased from day 10. AHH activities of the 3 and 10 mg/L treatment groups became stable from day 6 until the end of the experiment (P < 0.05). 3.2. Internal B[a]P dose and tissue lipid content B[a]P concentrations in gills and digestive glands were analyzed separately, and only the scallops of the seawater group were analyzed as control. Data concerning B[a]P dose in the

The results of mean F values of each individual treatment group are presented in Fig. 2. No significant difference in F values was observed over the 20-day exposure period for the control and acetone control groups (P > 0.05). F values of the control and acetone control groups, respectively, were 0.842  0.024 and 0.835  0.018 in gills, and 0.841  0.029 and 0.833  0.015 in digestive glands. F values of gill samples from the 3 and 10 mg/L treatment groups showed marked decreases (0.718  0.019 and 0.509  0.089, respectively) (P < 0.05) after 12 h of exposure, while the 0.5 mg/L B[a]P treatment groups showed marked decreases in mean F values (0.608  0.049, P < 0.05) after 1 day of exposure. The F values remained low and significantly different from the control values (P < 0.05) until day 6 for the 0.5 mg/L (0.246  0.082) and 3 mg/L (0.289  0.054) B[a]P

Table 1 B[a]P concentration (mean  SD) in gills and digestive glands of exposed and unexposed scallops Tissue

B[a]P dose (mg/L)

B[a]P concentration (mg/g w.w.) 0 days

3 days

10 days

20 days

Gills

Control 0.5 3 10

0.00146  0.0002a

0.00145  0.0002a 0.04742  0.0041b 0.17234  0.0147b 0.60853  0.0420b

0.00148  0.0002a 0.06814  0.0084c 0.24766  0.0319bc 0.97123  0.0687c

0.00147  0.0002a 0.08176  0.0113c 0.20983  0.0202c 0.99459  0.0284c

Digestive glands

Control 0.5 3 10

0.00512  0.0028a

0.00525  0.0040a 0.10361  0.0213b 0.45216  0.0191b 1.85887  0.0256b

0.00613  0.0018a 0.12605  0.0145b 0.59396  0.0198c 2.29784  0.0675c

0.00434  0.0019a 0.13890  0.0074b 0.57918  0.0310c 2.25898  0.0292c

Posthoc testing was performed using time as the factor, and B[a]P exposed concentration as the dependent list. The same superscript letters in the same row of the table mean insignificant difference ( p > 0.05), while different letters mean significant difference ( p < 0.05).

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Gills

1

F value

0.8 0.6 0.4 0.2 0

0

10

5

15

20

Time (days) Control

Acetone

0.5 µg/L

3 µg/L

10 µg/L

Digestive glands

1

F value

0.8 0.6 0.4 0.2 0

0

5

10

15

20

Time (days) Control

Acetone

0.5 µg/L

3 µg/L

10 µg/L

Fig. 2. Variations in the ratio of double-stranded DNA to total DNA (F value) (mean  SD) between scallop groups exposed to various concentrations of B[a]P over a 20-day exposure period (n ¼ 3).

treatment groups and then showed a gradual increase. At 20 days, the mean F values of the 0.5 mg/L treatment group (0.725  0.070) returned to the same level as the control (P > 0.05). The F values in the 10 mg/L treatment group decreased and remained stable after 3 days of exposure (P > 0.05). Digestive gland samples from these three treatment groups all showed marked decreases in mean F values (0.632  0.022, 0.596  0.026, 0.454  0.039, respectively, in the 0.5, 3 and 10 mg/L treatment groups) after 12 h of exposure (P < 0.05). The F value became lowest at day 6 for the 0.5 and 3 mg/L treatment groups, followed by a gradual increase in mean F values, and there was no significant difference in F values between the control and the 0.5 mg/L treatment group (0.691  0.041) after 15 days. The F values of the 3 and 10 mg/L treatment groups remained low and significantly different from the control values (P < 0.05) until the end of the experiment. 4. Discussion 4.1. Time and doseeresponse relationship of B[a]P exposure and AHH activity AHH, a component of the MFO system, is responsible for the initial oxidative step in the metabolism of a variety of

PAHs. AHH activity, according to the measurement method used reflects the metabolism of B[a]P. Previous studies had reported that AHH activity induced in Mytilus galloprovincialis and Modiolus modiolus collected from highly PAHs contaminated locations compared to that of mussels from a clean site (McDonald, 1990; Garrigues et al., 1990; Michel et al., 1994; Willett et al., 1999). In previous research, B[a]P metabolites in the gill and digestive gland of M. galloprovincialis were measured at 2, 8, 24, 48 h after injection with 3H-B[a]P and the results demonstrated that the mussel is able to develop a well balanced phase I and phase II xenobiotic metabolism limiting the accumulation of reactive metabolites (Michel et al., 1995). These published results appeared to indicate that AHH activity can be induced by PAHs in mussels. However, in the studies conducted with North Sea mussels injected with 5 mg/kg B[a]P or 20 mg/kg TCDD, no changes in AHH activities were observed 48 h after treatment (Willett et al., 1999). Our results are consistent with the field studies, AHH activities were induced in gills and digestive glands of the scallop C. farreri, and tissue differences in AHH activities were observed. AHH activity in gills of the 0.5, 3 and 10 mg/L treatment groups was induced at the 12th hour, and in digestive glands it was induced at day 3, day 1 and the 12th hour, respectively. AHH activity of the 0.5 and 3 mg/L treatment groups was induced earlier in gills than in digestive glands, which can be a consequence of gills contacting water directly and thereby being more sensitive to the xenobiotic. However, AHH activities in the digestive gland are consistently higher than in gills, for the function of this gland parallels that of the vertebrate liver (Livingstone, 1992); the results emphasized the different functionality of the two analyzed tissues. The AHH activity was constant at the latter period of the experiments, implying a dynamic metabolism balance. The induction mechanism of AHH activity is an AhR (aryl hydrocarbon receptor or dioxin receptor) mediated transcriptional activation, which has been reported in vertebrates (Safe, 1988, 1995; Junsei and Yoshiaki, 2003). But it has not been identified in marine or aquatic invertebrates. There are several studies on AhR or PAHs binding proteins; however, no determined result has been concluded in regulation of xenobiotic metabolizing enzyme activities in bivalves (Hahn et al., 1994; Willett et al., 1999; Butler et al., 2001). According to our results, the AHH activity was induced by B[a]P exposure, and the time-course and dose-dependent relationship were described, suggesting that there seems to be an induction mechanism in this scallop, and further studies are needed. 4.2. B[a]P accumulation The B[a]P tissue concentration results show that B[a]P was rapidly taken up in 3 days. B[a]P accumulation in the gill and digestive gland rose parallel with the exposure dose. These results are comparable to the studies in mussels (Mytilus sp.) reported by Canova et al. (1998) and Skarphe´ðinsdo´ttir et al. (2003). Time-course of B[a]P accumulation was shown in the results, and the time-course suggested an incoming plateau of B[a]P accumulation, implying a dynamic balance of B[a]P

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uptake and B[a]P metabolism. This was consistent with the result of the AHH enzyme experiment, which indicated that AHH activity was induced and became constant gradually until the end. B[a]P accumulation in the digestive gland was higher than that in gills, which is in accordance with what has been seen in other studies of B[a]P accumulation in mussels (Canova et al., 1998; Skarphe´ðinsdo´ttir et al., 2003). Tissue accumulation difference is determined by lipid levels because B[a]P is a hydrophobic substance. In our results, the lipid level in the digestive gland was about fourfold higher than that in the gill. Meanwhile, microstructure of the digestive gland of the scallop C. farreri showed food particles in digestive cells (Sheng et al., 2001), which indicated that the function and character of the digestive gland is also the reason for the highest B[a]P accumulation in this tissue. 4.3. DNA strand break Once taken up by organisms, PAHs can be metabolized to reactive intermediates through the action of MFO or peroxidative reaction (Stegeman, 1981; Mitchelmore et al., 1998; Pan et al., 2005). Monoxygenation and one-electron oxidation are currently considered the two main reactions for the activation of B[a]P and PAHs in general. Michel et al. (1995) found that after treatment of M. galloprovincialis with 3H-B[a]P injection, several polar metabolites including 9-10, 4-5 and 7-8 B[a]P diols (17%), 1e6, 3e6, and 6e12 B[a]P quinones (47%) and 9, 1, 3-B[a]P phenols (36%) were detected in the digestive gland and gill (Michel et al., 1995). It is worth noting that at least half the B[a]P metabolites in mussels are benzo[a]pyrene quinines (BPQs) which can undergo redox cycling to generate the superoxide anion radical (O) and other reactive oxygen species (ROS) (Martinez and Livingstone, 1995; Canova et al., 1998). Some of these reactive intermediates such as BPQs can directly bind to DNA forming DNA adducts, while others, comprising free radicals and ROS cause oxidative damage of DNA (Shugart, 1999). Ching et al. (2001) studied the B[a]P doseeresponse and time-course of DNA adduct formation and DNA strand breaks in the digestive gland of the mussel Perna viridis. In their results, mean F values of the 0.3 and 3 mg/L B[a]P treatment groups reached lowest twice, for the first time at day 1 and the second time at day 18 in 24 days’ exposure. In our results, there was another pattern of the time-course of F value. The F value in the gill and digestive gland of the 0.5 and 3 mg/L B[a]P treatment groups decreased until day 6 and then showed an gradual increase trend until the end of the experiment, while the F value in both tissues of the 10 mg/L B[a]P treatment group showed a decrease until day 3 and became constant gradually. Considering that the AHH activity and B[a]P accumulation showed a plateau at the end of the experiment, it is implied that the reactive intermediates may be maintained at a comparative persistent level, initiating the DNA repair system. The authors presume that DNA damage in the 0.5 and 3 mg/L B[a]P groups could be repaired as the repair system was initiated, while in the 10 mg/L B[a]P group the repair systems were overwhelmed. On the other hand, though in digestive glands the

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AHH activity is higher than in gills, the function and the high lipid content of this gland make it have a higher B[a]P concentration, while DNA strand break of the two tissues showed no significant difference. Canova et al. (1998) suggested that the underlying phenomena, such as differences in rates of detoxification enzymes as well as DNA repair system between the two tissues, could probably affect the results of DNA damage. Therefore, future studies of the DNA repair system are necessary. 5. Conclusion The results presented here confirm that B[a]P was metabolized and accumulated in the scallop C. farreri, subsequently resulting in DNA damage. The time-courses of the AHH activity, tissue concentration and DNA strand break reflect the toxicity mechanism of B[a]P, while the dose-dependent relationships of these parameters indicate that the combination of biological effects and tissue contaminant data can reflect the degree of pollution and the toxic effects, suggesting that the battery of parameters of the scallop C. farreri can be reasonable biomarkers for pollution monitoring. It is worth noting that tissue doses of B[a]P are detectable even in control groups. Considering that the scallop C. farreri is a commercial species, it is important for food security to control the water quality especially the organic pollutants. Acknowledgements This work was supported by the Science Foundation of The Key Laboratory of Mariculture, Ministry of Education, Ocean University of China. We thank the staff at the Laboratory of Physiology for help with sampling and taking care of the scallops. References Anderson, J.W., Jones, J.M., Steinert, S., Sanders, B., Means, J., et al., 1999. Correlation of CYP1A1 induction, as measured by the P450 RGS biomarker assay, with high molecular weight PAHs in mussels deployed at various sites in San Diego Bay in 1993 and 1995. Marine Environmental Research 48, 389e405. Bligh, E.G., Dyer, W.J., 1959. A rapid method of total lipid extraction and purification. Canadian Journal of Biochemistry and Physiology 37, 911e917. Bradford, M.M., 1976. A rapid and sensitive method for the quantitation of microgram quantities of protein utilizing the principle of protein dye binding. Analytical Biochemistry 72, 248e254. Butler, R.A., Kelley, M.L., Powell, W.H., Hahn, M.E., Van Beneden, R.J., 2001. An aryl hydrocarbon receptor (AHR) homologue from the soft-shell clam, Mya arenaria: evidence that invertebrate AHR homologues lack 2,3,7,8-tetrachlorodibenzo-p-dioxin and b-naphthoflavone binding. Gene 278, 223e234. Canova, S., Degan, P., Peters, L.D., Livingstone, D.R., Voltan, R., Venier, P., 1998. Tissue dose, DNA adducts, oxidative DNA damage and CYP1A-immunopositive proteins in mussels exposed to waterborne benzo[a]pyrene. Mutation Research 399, 17e30. Cavalieri, E.L., Rogan, E.G., 1995. Central role of radical cations in metabolic activation of polycyclic aromatic hydrocarbons. Xenobiotica 25 (7), 677e688.

198

L. Pan et al. / Environmental Pollution 153 (2008) 192e198

Cesarone, D.F., Bologenesi, C., Santi, L., 1979. Improved microfluorometric DNA determination in biological material using 33258 Hoechst. Analytical Biochemistry 100, 188e197. Chaty, S., Rodius, F., Vasseur, P., 2004. A comparative study of the expression of CYP1A and CYP4 genes in aquatic in vertebrate (fresh water mussel, Unio tumidus) and vertebrate (rainbow trout, Oncorhynchus mykiss). Aquatic Toxicology 69, 81e93. Ching, E.W.K., Siu, W.H.L., Lam, P.K.S., et al., 2001. DNA adduct formation and DNA strand breaks in green-lipped mussels (Perna viridis) exposed to benzo[a]pyrene: dose- and time-dependent relationships. Marine Pollution Bulletin 42, 603e610. Collier, T.K., Varanasi, U., 1991. Hepatic activities of xenobiotic metabolizing enzymes and biliary levels of xenobiotics in English sole (Parophrys vetulus) exposed to environmental contaminants. Archives of Environmental Contamination and Toxicology 20, 462e473. Daniel, F.B., Haas, D.L., Santi, L., 1985. Quantitation of chemically induced DNA strand breaks in human cells via an alkaline unwinding assay [J]. Analytical Biochemistry 144, 390e402. Everaarts, J.M., Sarkar, A., 1996. DNA damage as a biomarker of marine pollution: strand breaks in seastars (Asterias rubens) from the North Sea. Water Science & Technology 34 (7), 157e162. Fan, Z.J., Cote, R.P., 1990. Coastal zone of PR China: management approaches and institutions. Marine Policy 6, 305e314. Garrigues, P., Raoux, C., Lemaire, P., Ribera, D., Mathieu, A., Narbonne, J.F., Lafaurie, M., 1990. In situ correlations between polycyclic aromatic hydrocarbons (PAH) and PAH metabolizing system activities in mussels and fish in the Mediterranean sea: preliminary results. International Journal of Environmental and Analytical Chemistry 38, 379e387. Goldberg, E.D., 1978. The surveillance of coastal marine waters with bivalves: the mussel watch. In: Albaige´s, J. (Eds.), Analytical Techniques in Environmental Chemistry: Proceedings of the International Congress, Barcelona, Spain, pp. 373e386. Hahn, M.E., Poland, A., Glover, E., Stegeman, J.J., 1994. Photoaffinity labeling of the Ah receptor: phylogenetic survey of diverse vertebrate and invertebrate species. Archives of Biochemistry and Biophysics 310, 218e228. Huggett, R.J., Kimerle, R.A., Mehrle, P.M., Bergman, H.L., 1992. Biomarkers: Biochemical, Physiological, and Histological Markers of Anthropogenic Stress. Lewis Publishers, Boca Raton, pp. 347. Junsei, M., Yoshiaki, F., 2003. Functional role of AhR in the expression of toxic effects by TCDD. Biochimica et Biophysica Acta 1619, 263e268. Kurelec, B., Krea, S., Lucic, D., 1996. Expression of multixenobiotic resistance mechanism in a marine mussel Mytilus galloprovincialis as a biomarker of exposure to polluted environments. Comparative Biochemistry and Physiology, Part C 113 (2), 283e289. Livingstone, D.R., 1992. Persistent pollutants in marine invertebrates. In: Walker, C.H., Livingstone, D.R. (Eds.), Persistent Pollutants in Marine Ecosystems. Pergamon, Oxford, UK, pp. 3e34. Mai, B.X., Fu, J.M., Zhang, G., Lin, Z., Min, Y.S., et al., 2001. Polycyclic aromatic hydrocarbons in sediments from the Pearl river and estuary, China: spatial and temporal distribution and sources. Applied Geochemistry 16, 1429e1445. Martinez, P.G., Livingstone, D.R., 1995. Benzo[a]pyreneedione-stimulated oxyradical production by microsomes of digestive gland of the common mussel, Mytilus edulis L. Marine Environmental Research 39, 185e189. McDonald, S.J., 1990. Benzo(a)pyrene Metabolism in a Deep Sea Mussel Associated with Natural Petroleum Seeps in the Gulf of Mexico. Ph.D. dissertation, Texas A&M University, College Station, TX, pp. 1e153. McElroy, A., Leitch, K., Fay, A., 2000. A survey of in vivo benzo[a]pyrene metabolism in small benthic marine invertebrates. Marine Environmental Research 50, 33e38. Menzie, C.A., Potocki, B.B., Santodonato, J., 1992. Exposure to carcinogenic polycyclic aromatic hydrocarbons in the environment. Environmental Science and Technology 26, 1278e1284.

Michel, X.R., Beasse, C., Narbonne, J.F., 1995. In vivo metabolism of benzo (a)pyrene in the mussel Mytilus galloprovincialis. Archives of Environmental Contamination and Toxicology 28, 215e222. Michel, X.R., Salaun, J.P., Galgani, F., Narbonne, J.F., 1994. Benzo(a)pyrene hydroxylase activity in the marine mussel Mytilus galloprovincialis: a potential marker of contamination by polycyclic aromatic hydrocarbon-type compounds. Marine Environmental Research 38, 257e273. Mitchelmore, C.L., Birmelin, C., Chipman, J.K., Livingstone, D.R., 1998. Evidence for cytochrome P-450 catalysis and free radical involvement in the production of DNA strand breaks by benzo[a]pyrene and nitroaromatics in mussel (Mytilus edulis L.) digestive gland cells. Aquatic Toxicology 41, 193e212. Pan, L.Q., Ren, J.Y., Liu, J., 2005. Effects of benzo(k)fluoranthene exposure on the biomarkers of scallop Chlamys farreri. Comparative Biochemistry and Physiology, Part C 141, 248e256. Quinn, B., Gagne´, F., Weber, J.-P., Blaise, C., 2005. Ecotoxicological effects of a semi-submerged municipal dump (Castle harbour, Bermuda) on the Calico scallop Argopecten gibbus. Marine Pollution Bulletin 51, 534e544. RNO, 1988. Re´seau National d’Observation de la qualite´ du milieu marin: Surveillance du milieu marin. IFREMER DERO/EL/RNO/56/MJ, 88. Safe, S.H., 1988. The aryl hydrocarbon (Ah) receptor. ISI Atlas of Science: Pharmacology 2, 78e83. Safe, S.H., 1995. Modulation of gene expression and endocrine response pathways by 2,3,7,8-tetrachlordibenzo-p-dioxin and related compounds. Pharmacological Therapy 67, 247e281. Schirmer, K., Chan, A.G.J., Greenberg, B.M., Dixon, D.G., Bols, N.C., 1998. Ability of 16 priority PAHs to be photocytotoxic to a cell line from the rainbow trout gill. Toxicology 127, 143e155. Sheng, X.Z., Liu, X.Y., Ren, S.L., et al., 2001. A histological and histochemical study on digestive diverticula of Farreris’s scallop, Chlamys (Azumapecten) farreri. Journal of Ocean University of Qingdao 31 (3), 361e367. Shugart, L.R., 1999. Structural damage to DNA in response to toxicant exposure. In: Forbes, V.E. (Ed.), Genetics and Ecotoxicology. Taylor & Francis, London, pp. 151e168. Siu, W.H.L., Hung, C.L.H., Wong, H.L., et al., 2003. Exposure and time dependent DNA strand breakage in hepatopancreas of green-lipped mussels (Perna viridis) exposed to Aroclor 1254, and mixtures of B[a]P and Aroclor 1254. Marine Pollution Bulletin 46, 1285e1293. Skarphe´ðinsdo´ttir, H., Ericson, G., Dalla, Z.L., Gilek, M., 2003. Tissue differences, doseeresponse relationship and persistence of DNA adducts in blue mussels (Mytilus edulis L.) exposed to benzo[a]pyrene. Aquatic Toxicology 62, 165e177. Stansbury, K.H., Flesher, J.W., Gupta, R.C., 1994. Mechanism of aralkyl-DNA adduct formation from benzo[a]pyrene in vivo. Chemical Research in Toxicology 7 (2), 254e259. Stegeman, J.J., 1981. Polynuclear aromatic hydrocarbons and their metabolism in the marine environment. In: Gelboin, H. (Ed.), Polycyclic Hydrocarbons and Cancer. Academic Press, New York, pp. 1e60. Van Veld, P.A., Westbrook, D.J., Woodin, B.R., Hale, R.C., Smith, C.L., Huggett, F.J., Stegeman, J.J., 1990. Induced cytochrome P450 in intestine and liver of spot (Leisostomus xanthurus) from a polycyclic aromatic hydrocarbon contaminated environment. Aquatic Toxicology 17, 119e132. Walker, C.H., Hopkins, S.P., Sibly, R.M., Peakall, D.B., 1997. Principles of Ecotoxicology. Taylor and Francis, London, pp. 1e321. Willett, K.L., Wilson, C., Thomsen, J., Porter, W., 1999. Evidence for and against the presence of polynuclear aromatic hydrocarbon and 2,3,7,8-tetrachloro-p-dioxin binding proteins in the marine mussels, Bathymodiolus and Modiolus modiolus. Aquatic Toxicology 48, 51e64.