Antioxidant responses to polycyclic aromatic hydrocarbons and organochlorine pesticides in green-lipped mussels (Perna viridis): Do mussels “integrate” biomarker responses?

Marine Pollution Bulletin 57 (2008) 503–514

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Marine Pollution Bulletin journal homepage: www.elsevier.com/locate/marpolbul

Antioxidant responses to polycyclic aromatic hydrocarbons and organochlorine pesticides in green-lipped mussels (Perna viridis): Do mussels ‘‘integrate” biomarker responses? Bruce J. Richardson *, Eva Mak, Sharon B. De Luca-Abbott, Michael Martin, Katherine McClellan, Paul K.S. Lam Department of Biology and Chemistry, Research Centre for Coastal Pollution and Conservation, City University of Hong Kong, 83 Tat Chee Avenue, Kowloon, Hong Kong

a r t i c l e

i n f o

Keywords: Polycyclic aromatic hydrocarbons (PAHs) Organochlorine pesticides (OCs) Perna viridis Antioxidant responses Body burden Dose regime

a b s t r a c t Polycyclic aromatic hydrocarbons (PAHs) and organochlorine pesticides (OCs) are generally present in the marine environment in complex mixtures. The ecotoxicological nature of contaminant interactions, however, is poorly understood, with most scientific observations derived from single contaminant exposure experiments. The objective of this experiment was to examine dose-response relationships between antioxidant parameters and body contaminant levels in mussels exposed to different exposure regimes under laboratory conditions. Accordingly, the green-lipped mussel, Perna viridis, was challenged with a mixture of PAHs (anthracene, fluoranthene, pyrene, benzo[a]pyrene) and OC pesticides (a-HCH, aldrin, dieldrin, p,p0 -DDT) over a 4 week period. Contaminants were delivered under four different dosing regimes, with all treatments receiving the same total contaminant load by the end of the exposure period. Antioxidant biomarkers were measured after 1, 2, 3 and 4 weeks, including glutathione (GSH), gluathione-S-transferase (GST), superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx), glutathione reductase (GR) and lipid peroxidase (LPO). GST and CAT were induced in hepatic tissues in most of the exposure regimes, with the majority of significant induction occurring in a constant exposure regime and a two-step alternate exposure regime. Significant differences among exposure regimes were detected in the body burden of contaminants after 28 days. Hepatic CAT and GSH are proposed as potentially useful biomarkers as they showed good correlation with target contaminants and were not readily affected by different dosing patterns. Ó 2008 Elsevier Ltd. All rights reserved.

1. Introduction Persistent organic pollutants (POPs) are carbon-based compounds that resist photochemical, biological and chemical degradation (Eduljee, 2001). POPs are typically characterised by low water solubility and high lipid solubility, which leads to persistence and bioaccumulation. In the marine environment, POPs partition strongly to solids and readily accumulate in fatty tissues of organisms. POPs can be deposited into the marine environment directly through atmospheric deposition, river inflow, discharges of municipal or industrial wastewater and by leaching from contaminated sediments and dump sites (Bernes, 1998). The marine environment in Hong Kong is under severe stress from various anthropogenic activities. Elevated concentrations of POPs, including PAHs and OCs, have been observed in waters, biota, and sediments since 1970 (Wu, 1988; Richardson et al., 2000; EPD,

* Corresponding author. Tel.: +852 2788 7042; fax: +852 2788 7406. E-mail address: [email protected] (B.J. Richardson). 0025-326X/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.marpolbul.2008.02.032

2000), and their potential as toxicologically significant contaminants has been assessed in recent studies of Hong Kong waters. PAHs are primarily formed from incomplete combustion of fossil fuels and other organic material, and they enter the marine environment from atmospheric deposition, direct discharge and runoff from land (Piccardo et al., 2001; Neff, 1979). Zheng and Richardson (1999) noted that in Hong Kong nearshore waters, concentrations were indicative of serious pollution in areas such as Victoria Harbour and Tolo Harbour. In a study of marine sediments in Victoria Harbour, Hong et al. (1995) determined the concentration of total PAHs to be between 1.2 and 454 lg g1 (d.w.). Furthermore, the Environmental Protection Department in Hong Kong began routine monitoring of PAHs in marine sediments in 1984, and determined that anthracene, fluoranthene and pyrene were generally the dominant components, accounting for approximately 40% (by dry weight) of the total PAHs (see Richardson et al., 2000 and references therein). Although Hong Kong has little arable land, pesticides (including HCHs, DDT and metabolites, dieldrin, aldrin and chlordane) are commonly found in coastal waters, sediment and biota (Richardson

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and Zheng, 1999), and it is likely that cross-border influences play a significant role in their distribution. OCs reach the marine environment primarily from surface runoff and ground leachate, but are also found in wastewater and stormwater discharges (Clendening et al., 1990). Zheng (2001) detected HCHs in the range 0–2.3 ng g1 dry weight (d.w.) and 1.38–30.3 ng g1 (d.w.) for DDT (and metabolites) in marine sediments. Furthermore, Hong et al. (1995) detected total DDT in very high levels in surface sediment from two typhoon shelters (56 and 97 ng g1 d.w.). Previous studies have shown the usefulness of mussels as bioindicators of chemical contaminants, and they continue to be used effectively to monitor pollution in the marine environment. Chemical analyses measure the contaminants present, but do not necessarily reveal potential biological effects. The need to detect and assess the effects of contaminants, especially at low concentrations in complex mixtures, has led to the development of molecular markers of contaminant effects (i.e. biomarkers). There has been varied success with the use of antioxidant responses in marine mussels as biomarkers of persistent organic pollutants (De Luca-Abbott et al., 2005; Cheung et al., 2002; Akcha et al., 2000; Eertman et al., 1995; Livingstone et al., 1990). Antioxidant responses are often induced by exposure to reactive oxygen species or contaminants. These defence systems remove chemically reactive intermediates produced by xenobiotic metabolism and endogenous pathways (Kappus, 1987) and are also involved in eliminating electrophilic chemicals or metabolites in enzymatic pathways. Some of the more commonly used antioxidant biomarkers include superoxide dismutase (SOD), catalase (CAT), glutathione peroxidase (GPx), glutathione transferase (GST), glutathione (GSH), glutathione reductase (GR), and lipid peroxidation (LPO). The SODs are a group of metalloenzymes that protect against oxidative damage by catalysing the dismutation of the superoxide anion radical to hydrogen peroxide (H2O2) plus water (Fridovich, 1995; Scandalios, 1993). SOD increased in Scapharca inaequivalvis when exposed to copper for 14 days (Isani et al., 2003). The primary role of CAT is to decompose H2O2, although some activity is associated with methyl- and ethyl-hydroperoxides. CAT is present in most eukaryotic cells, and the enzyme is primarily localised within the peroxisomes (Akcha et al., 2000; Khessiba et al., 2001). Increased CAT activity has been observed in ribbed mussels exposed to paraquat (Wenning and Di Giulio, 1988; Wenning et al., 1988). An alternative route for hydrogen peroxide reduction is the glutathione dependent pathway. Glutathione is an essential component of antioxidant defences in eukaryotic cells. GPx removes H2O2 by coupling its reduction to H2O with oxidation of reduced GSH. Compounds with electrophilic centres spontaneously conjugate with GSH or via enzymatic catalysis by GST. According to Commandeur et al. (1995), GST reacts with xenobiotics by: (1) forming direct-acting GSH conjugates; (2) acting as a transporter molecule that releases reversibly bound electrophiles at specific tissues; (3) forming GST conjugates that are bioactivated by subsequent metabolism of the GSH moiety; and (4) performing reductive bioactivation mechanisms. GSH is an important scavenger against oxidative stress (Allen and Sohal, 1986), and induction of GSH reflects an adaptation to pollutants and has been shown to play a critical role in maintaining cellular homeostasis (Doyotte et al., 1997). GR is important in maintaining a critical ratio of GSH/GSSG (gluathione disulfide) under oxidative stress. Recent studies, using mussels, have detected a marked increase in GR upon exposure to PAHs (Gamble et al., 1995; Akcha et al., 2000; Porte et al., 1991, 2001). Lipid peroxidation is a complex process in which unsaturated lipid material reacts with molecular oxygen to yield lipid hydroperoxides, which are degraded to a variety of products including alkanals, alkenals, hydroxyalkenals, ketones and alkanes

(Mead, 1976). Consequences of membrane lipid peroxidation include the perturbation of various cellular and organelle membrane functions and damage to DNA and proteins. Levels of chemical pollutants in the environment often display temporal and seasonal variation. For example, Boryslawsky et al. (1985) reported marked temporal changes over short periods in the concentration of dieldrin in water in North England. The use of biomarkers, such as antioxidant response, in biomonitoring studies is complicated by this variation. Such fluctuations will affect the performance of biomarkers in the field and complicate data interpretation. Therefore, four exposure regimes were established in the present study in order to investigate the effects of dose variation on antioxidant response and uptake of pollutants in the green-lipped mussel, Perna viridis, during a 28-day experiment.

2. Materials and methods Five hundred green-lipped mussels, P. viridis (L.), were collected from a fish farm at Sha Tau Kok, north-eastern Hong Kong, and transported to the laboratory where they were acclimated in clean seawater for 10 days. Seawater was exchanged every 24 h. As a food source for filter-feeding molluscs, phytoplankton may facilitate the uptake of water-insoluble contaminants into organisms, increasing the potential for toxicity. To control this factor as much as possible, constant amounts of algae were supplied daily during the experiment. Unicellular green algae (Dunaliella tertiolecta), at an initial density of 2  106 cell L1 of seawater, were supplied daily to feed mussels during the acclimation and exposure periods. A preliminary study of contaminant concentrations in mussels exposed at ‘‘control” and ‘‘polluted” sites in Hong Kong was conducted using mussels (P. viridis) and semi-permeable membrane devices (SPMDs) in order to determine the environmentally realistic level of target contaminants in the field (see Richardson et al., 2001, 2005; Prest et al., 1995). The level of target contaminants in water was back-calculated using a Kow approach. Our laboratory experiment subsequently comprised four contaminant treatments (each with two replicate tanks); individual treatments received a different dose of PAHs and OCs each week. Upon completion of the 4-week exposure period, each treatment had received the same total dose of contaminants. A mixture of PAHs (anthracene, fluoranthene, pyrene, B[a]P) and OC pesticides (a-HCH, aldrin, dieldrin, p,p0 -DDT) was prepared in acetone and used in the experiment; nominal weekly exposure regimes are shown in Table 1. In addition, there were two replicate solvent control treatments (receiving acetone only) and two control (no solvent) treatments. The dosing regimes, and the rationale for their usage, were similar to those utilised by us in previous experiments (e.g. Siu et al., 2004). Twenty mussels (P. viridis) with shell lengths 90–110 mm were placed in each tank with 50 L of clean seawater. Contaminants (as appropriate) were added to the freshly replaced seawater in each aquarium on a daily basis. Four mussels were harvested from each tank every week (days 7, 14, 21, 28) and control mussels were collected prior to the commencement of the experiment (Day 0). The amount of seawater in the aquaria was adjusted proportionally after mussels were removed each week; i.e., after the day 7 samples were collected, each tank contained 16 mussels and only 40 L of water, after day 14 each tank held 12 mussels and 30 L, etc. Mussel flesh, and gill and hepatopancreas tissues were carefully dissected, weighed and stored separately at 20 °C for later analysis (see Cheung et al., 2001, 2002, 2004 for details and rationale for the use of gill and hepatopancreas tissues for antioxidant measurements). Physicochemical measurements were performed daily on both freshly spiked seawater and 1-day spent seawater, and no signifi-

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B.J. Richardson et al. / Marine Pollution Bulletin 57 (2008) 503–514 Table 1 Daily exposure concentrations (lg L1 day

1

) of each PAH and OC pesticide for the various exposure regimes

Exposure regime

A. Constant B. 2-step alternate pulse (begin with exposure) C. 2-step alternate pulse (begin without exposure) D. 4-step alternate pulse (begin with exposure) E. 4-step alternate pulse (begin without exposure) F. 4-step increasing G. 4-step decreasing H. Solvent control I. Control (without solvent)

PAHs

OCs

Week 1

Week 2

Week 3

Week 4

Week 1

Week 2

Week 3

Week 4

1 2 0 2 0 0.7 1.3 0 0

1 2 0 0 2 0.9 1.1 0 0

1 0 2 2 0 1.1 0.9 0 0

1 0 2 0 2 1.3 0.7 0 0

0.05 0.1 0 0.1 0 0.02 0.08 0 0

0.05 0.1 0 0 0.1 0.04 0.06 0 0

0.05 0 0.1 0.1 0 0.06 0.04 0 0

0.05 0 0.1 0 0.1 0.08 0.02 0 0

cant discrepancies among aquaria were detected. Average temperature, dissolved oxygen, salinity and pH were 22.3 ± 1.3 °C, 6.1 ± 1.2 mg/L, 31.7 ± 0.9 ‰, and 7.7 ± 0.3 respectively. Any dead mussels (valves widely opened and not responding to external stimulation) were removed from the aquaria immediately and replaced with depurated mussels of a similar size in order to maintain a consistent density. Replacement mussels were appropriately tagged to ensure they were not used for body burden or physiological response. 2.1. Antioxidant response assays Gill and hepatopancreas tissue samples were thawed on ice and homogenized in buffer (50 mM potassium phosphate buffer, 0.1 M KCl, 0.1 mM EDTA, pH 7.4; with 20% glycerol to protect the enzymes) using an Ultra-Turrax tissue homogenizer. The homogenate was centrifuged for 30 minutes at 12,000g and 4 °C. The supernatant (post-mitochrondrial fraction) was separated into five equal portions of 200 ll, placed in pre-cooled microtubes and stored at 80 °C for further biochemical analysis (Birmelin et al., 1998). The protein content of each sample was determined using a BioRad Protein Assay Kit. All enzyme determinations were carried out in triplicate at 25 °C and were based on methods of Cheung et al. (2004). GST activity was determined using 1-chloro-2,4-dinitrobenzene (CDNB) as the substrate (Jakoby, 1985). One unit of GST was defined as the amount of glutathione conjugate formed using 1 mM GSH and CDNB/min per mg protein. GSH was quantified using a modification of the method described by Anderson (1998). The assay was based on the enzymatic GSSG reductase assay, in which the reaction mixture contained phosphate buffer, NADPH, dithionitrobenzoic acid and GSSG reductase. The reaction was monitored at 412 nm and the amount of GSH was expressed as nmol GSH/ mg protein. The SOD assay was based on the ability of SOD to inhibit the auto-oxidation of pyrogallol in a 50 mM Tris-succinate buffer (pH 8.2) (Marklund and Marklund, 1974). The rate of pyrogallol oxidation was monitored spectrophotometrically at 420 nm and one unit of SOD was defined as the capability to cause 50% inhibition of the pyrogallol oxidation process. The method for determining GR activity was based on that of Carlberg and Mannerrik (1985) which followed the oxidation of 2 mM NADPH by GR in the presence of 20 mM oxidized glutathione (GSSG) at 340 nm. GPx was measured in a coupled enzyme system where the GSSG formed in the GPx reaction was converted to the reduced form by glutathione reductase (Lawrence and Burk, 1976). The consumption of NADPH was monitored at 340 nm. Degradation of H2O2 by CAT was monitored colorimetrically using ferrous sulphate and potassium thiocyanate (Cohen et al., 1996) where the remaining H2O2 subsequently oxidizes ferrous ions to ferric ions and forms a red coloured ferrithiocyanate with thiocyanate ions (measured at 492 nm). One unit of CAT was defined as [In(A1/A2)/(t1  t2)]/mg protein, where A1 is the absorbance TM

at time point 1 (t1), and subsequent time periods. Lipid peroxidation was quantified as thiobarbituric acid reactive substances. The aldehyde formed was estimated spectrophotometrically at 535 nm. Lipid peroxidation was expressed as nmol of MDA formed per mg protein. 2.2. Chemical analyses Methods used to extract the PAHs and OCs have been described by Xu et al. (1999) and Cheung et al. (2004). In summary, the remaining soft tissue of each mussel was freeze-dried for 24 h and then homogenized to a powder using a tissue grinder. The powder was then extracted in dichloromethane and anhydrous sodium sulphate and the extract centrifuged at 3000 rpm for 10 min. The extraction and centrifugation process was repeated twice, all three supernatants were combined and the volume reduced to approximately 5 mL by rotary evaporation under reduced pressure. A silica gel fractionation column (internal diameter 1 cm with fused-in filter disc) was used for sample clean-up. The column contained 13 cm3 of activated silica gel (average pore size: 60 Å). After loading the sample, the column was successively eluted with 15 ml hexane and 30 ml of 20% dichloromethane in hexane. The elution volume was reduced to approximately 1 ml by nitrogen blowdown before analysis. PAHs and OC pesticides were analyzed using a Hewlett Packard 6890 gas chromatograph equipped with a FID and ECD. A 2 lL sample and external standard were injected automatically by a HP 7673 automatic controller. A HP-5MS column (Hewlett Packard; 30 m  0.32mm  0.25 lm) was used for separating target contaminants. External standards of the PAH and OC compounds were used to calculate response factors by comparing their peak areas with an internal standard (m-terphenyl). Peaks were identified by retention time. Lipid content of the mussel extract was determined by drying the filtrate using nitrogen blow-down until a constant weight was reached. All data were expressed as ng g1 lipid weight (Pruell et al., 1987). Recoveries of PAHs and OCs using this method were 82–102% and 78–103%, respectively. Coefficients of variation no greater than 9% were obtained for the range of compounds. Data were not corrected for recovery. 2.3. Statistical analyses Statistical differences between control and treatment antioxidant responses under different exposure regimes were tested using Mann–Whitney tests. Two-way analyses of variance (ANOVA) on ranked data were used to identify differences in enzyme activity by exposure regime and exposure time. Tukey’s multiple comparison tests were used where significant differences were detected in the ANOVA. Tissue contaminant levels were compared with log Kow values using linear regression models. Data points that were below the detection limit were replaced by the value of the detection limit

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of the GC system. Body burden data were compared among exposure regimes using ANOVAs. Statistical relationships between individual body contaminant concentrations and antioxidant parameters in both tissue types were compared using non-parametric Spearman’s rank correlation. 3. Results 3.1. Antioxidant parameter responses Induction of hepatic and gill antioxidant parameters was observed in different exposure regimes. Of the target antioxidant parameters, hepatic GSH, CAT and GPx levels were significantly induced in most of the exposure regimes compared to the control (Table 2). Hepatic CAT activity was significantly induced in regimes A and B in the first week of exposure, whereas induction occurred in the third and fourth week in regime C, and only in the fourth week in regime F (Table 2). GSH levels in hepatic tissue significantly increased in all exposure regimes, except D and F. Induction of GSH was observed in the four consecutive sampling weeks in regime B and the first week in regime A. GSH levels significantly increased in week two and three in regime E, but only in week 3 in regime C and week 4 in regime G (Table 2). GPx activity in hepatic tissue was significantly higher in regime B in week one and two, in regime E in week two and in regime G in weeks one and three. There was minimal induction of SOD, GR, GST and LPO during the exposure period, regardless of regime. GST and SOD were, however, significantly higher than the control in weeks three and four in regime B (Table 2).

a

50

Antioxidant responses in gill tissue were markedly less than that observed in hepatic tissue (Table 2). LPO was significantly induced in regime A in the first sampling week, and this level was maintained throughout the sampling. CAT was significantly higher than the control in the first week in regimes A and C, and in the third week in regime E. There was no significant induction of gill GPx during the experiment, and GST and GSH were only induced on a single sampling occasion in one regime. An increase in gill SOD activity was detected in regimes B, D and E in week 2, and in week 4 in regime B. GR was significantly higher than that of the control in regime C in weeks 1 and 4, in week 3 in regime F and in week 1 in regime G (Table 2). After four weeks exposure, hepatic antioxidant responses were generally higher than that in gill, excluding lipid peroxidation which was detected at higher levels in gill tissues. GSH levels, CAT, SOD and GST activities in hepatopancreas were significantly higher than the control group after 28 days exposure in regime B (see Table 2). Similarly, at the end of the exposure period, CAT was higher in regime C and F, whereas gill lipid peroxidation was significantly higher in regime A only. In regime G, only hepatic GSH levels were statistically higher than the control after 28 days. No significant differences in antioxidant parameters were detected after 28 days in regimes A, D and E, and there was no difference in GR or GPx activity at the end of the exposure period, regardless of regime (except for gill GR in regime C). Over the 28 days, there were significant differences in antioxidant responses in gill tissue in regimes A, B and C. Gill SOD activity and lipid peroxidation were higher than the control at the completion of the experiment in regimes B and A respectively, whereas hepatic GR in regime B was significantly lower than the control.

b

Tissue PAHs

25

40

20

20 10

0

1

2

Anthracene

3

Fluoranthene

Pyrene

4 B(a)P

20

15

10

5

Target Contaminants in Mussel Tissue (ng g- 1 lipid weight)

Target Contaminants in Mussel Tissue (ng g-1 lipid weight)

30

0

30

15 10 5 0

0

1

2

Anthracene

3

Fluoranthene

4

Pyrene

B(a)P

60 50 40 30 20 10

0

0

0

1

2 HCH

3

Aldrin

Dieldrin DDT

Weeks

4

0

1

2 HCH

Aldrin

3 Dieldrin

4 DDT

Weeks

Fig. 1. Concentrations of individual PAHs and OC pesticides in mussels exposed to (a) regime A; (b) regime B; (c) regime C; and (d) regime D. For details of dosing in each regime, see Table 1.

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c

d

100

40

80 30

Target Contaminants in Mussel Tissue (ng g-1 lipid weight)

Target Contaminants in Mussel Tissue (ng g-1 lipid weight)

60 40 20 0 0

1 Anthracene

2

3

Fluoranthene

Pyrene

4 B(a)P

100 80 60 40

20

10

0 0

1

2

Anthracene

3

Fluoranthene

4

Pyrene

B(a)P

100 80 60 40 20

20

0

0 0

1 HCH

2

Aldrin

3

Dieldrin

4

0

1

2 HCH

DDT

3

Aldrin

Dieldrin

4

DDT

Weeks

Weeks Fig. 1 (continued)

3.2. Tissue contaminant uptake Contaminant body burden was analysed in mussels from the regimes where most antioxidant activity was detected (A, B, C, D and controls). Body burden of contaminants in solvent control and control mussels did not significantly increase from low levels throughout the exposure period. In regime A, the concentration of PAHs, excluding anthracene, increased significantly with increased exposure time (Fig. 1a). Regression models confirmed there was a linear increase in concentration of B[a]P (r2 = 0.950), fluoranthene (r2 = 0.958) and pyrene (r2 = 0.958). Anthracene, however, did not show this pattern and remained at low tissue concentrations throughout the exposure period. A higher level of dieldrin and p,p0 -DDT was observed in weeks 1 and 2 in regime A, but concentrations declined in the subsequent weeks (Fig. 1a). There was no increase in the body tissue levels of HCH and aldrin in any of the experimental treatments. The concentrations of PAHs and two of the OC pesticides mirrored the nominal concentrations in water in regime B. Body concentrations increased rapidly in the first two weeks, and were depleted in weeks 3 and 4 (Fig. 1b). Of the OCs, only dieldrin and p,p0 -DDT were accumulated to high levels in mussel tissue, and they showed the same pattern as the PAHs with levels increasing over the first two weeks and marginally declining in weeks three and four (Fig. 1b). The OC pesticides were retained in relatively higher concentrations (percentage basis) than PAHs in mussel tissues. In regime C, the concentrations of PAHs and OCs in mussel tissues were closely correlated with the contaminant dosing concen-

trations and showed a pattern of exponential increase over time (Fig. 1c). As with regime A and B, however, levels of HCH and aldrin did not increase during the exposure period. Correlation analysis gave the following results: B[a]P r2 = 0.900; fluoranthene r2 = 0.991; pyrene r2 = 0.929; anthracene r2 = 0.647; p,p0 -DDT r2 = 0.841; and dieldrin r2 = 0.853. PAH and OC concentrations in mussel tissue closely followed the dosing concentration in regime D (Fig. 1d). Similarly, as with regime B, PAHs were more readily depurated (on a percentage basis) than the OC pesticides in the weeks where no dose of contaminants was delivered. The mussel body burden of PAHs was well correlated with the log Kow values (Table 3) of the respective compounds in regime A (r2 = 0.9014). The sum of the concentrations of anthracene, fluoranthene and pyrene was well correlated with log Kow values in regime B (r2 = 0.638), C (r2 = 0.996) and D (r2 = 0.9693), respectively. Interestingly, B[a]P was not well correlated with Kow, as was the case for OC pesticides in all regimes. According to the experimental design, the total amount of individual contaminants delivered to each regime was the same after 28 days. Different final concentrations of PAHs and OC pesticides were, however, detected among regimes (Fig. 2). Body burdens of total PAHs were higher in regime C compared to regimes A, B, D and the control, whereas OC pesticide concentrations were higher in regime C compared to A, B and the control. No significant difference was detected in total OC levels between regime C and D. Total PAHs were higher in regime A compared to regime B and D, whereas total OCs were higher in regime B compared to regime A (Fig. 2).

B.J. Richardson et al. / Marine Pollution Bulletin 57 (2008) 503–514

Table 2 Hepatic (*) and gill (#) antioxidative responses in mussels exposed to different contaminant dose regimes A - G (see Table 1), showing significant differences compared with control (*/# P < 0.05, **/## P < 0.01, ***/### P < 0.001). Dosing Regime

CAT Week1 Week 2 Week 3 Week 4 GSH Week1 Week 2 Week 3 Week 4 GPx Week1 Week 2 Week 3 Week 4 GST Week1 Week 2 Week 3 Week 4 SOD Week1 Week 2 Week 3 Week 4 GR Week1 Week 2 Week 3 Week 4 LPO Week1 Week 2 Week 3 Week 4

A

B

C

*** ##

***

#

*** ***

*** ***

*** *

*** *** *** ***

D

E

#

F

G

***

150 100 50

A ***

C Total PAH

*

** ***

B

D

Control

Regimes

* *

*

Total OCs

Fig. 2. Body burdens of PAHs and OC pesticides after 28 days exposure.

** *

###

* *

## *** ** #

*

*

#

#

#

# ##

## # # # #

**

positive correlation), as well as total PAHs and total OC pesticides (Table 5). Similarly, a negative correlation was detected between gill LPO and individual OCs, as well as total OCs. No relationship was detected between PAHs (except pyrene) and gill LPO. Gill GR activity was positively correlated with dieldrin, p,p0 -DDT and total OC pesticides, but was not correlated with each of the tissue PAHs, nor with total PAHs. No consistent relationship was observed between gill SOD, GST, GSH and CAT with individual PAHs and OC pesticides. Positive correlations were only observed between gill SOD and B[a]P. Gill SOD was, however, negatively correlated with HCH. Gill GSH was positively correlated with only B[a]P and gill CAT was negatively correlated with pyrene, B[a]P, HCH and aldrin. Gill GST showed no correlation with any target contaminant (Table 5). Correlations between each antioxidant parameter in hepatic tissue revealed positive associations between GPx and GSH, GR and CAT, SOD and GSH, SOD and CAT, and LPO and CAT. In addition, there was a negative relationship between GPx and LPO in hepatic tissue. In gill tissue, significant positive correlations were detected between GPx and GSH, GR and CAT, and LPO and CAT. 4. Discussion

PAHs

LogKow

OC pesticides

LogKow

Anthracene Fluoranthene Pyrene B(a)P

4.54a 5.22a 5.13a 6.50a

a-HCH Aldrin Dieldrin p,p’-DDT

3.89b 5.66b 5.16b 5.98b

a

200

0

Table 3 Log Kow values of target contaminants

b

250

Body Burden (ng/g lipid) weight)

508

PAH data from Mackay et al. (1980). Pesticides data from Isnard and Lambert (1988).

3.3. Relationship between antioxidant responses and target contaminants Positive correlations between hepatic GSH level and each of the PAHs and OC pesticides were observed (Table 4). Total PAHs and total OCs were also positively correlated with hepatic GSH response. CAT activity in hepatic tissue was positively correlated with anthracene, pyrene, each OC pesticide, total PAHs and total OC pesticides. A positive correlation was detected between hepatic SOD activity and each individual PAH, total PAHs, HCH and p,p0 DDT. Hepatic GR was positively correlated with individual OCs, total OCs and anthracene (Table 4). No correlations were detected between contaminant concentrations and LPO, and no consistent relationship was observed between contaminant concentrations and GST and GPx (Table 4). Negative correlations were observed between gill GPx activity and each individual contaminant (except aldrin was a slight but

Recent studies have focused on the utility of antioxidants as biomarkers for marine invertebrates (Livingstone, 1991a,b,1993; Livingstone et al., 1989; Livingstone et al., 1992; Porte et al., 1991; Sole et al., 1996). Kappus (1986) revealed that antioxidant defences consist of three classes that include water-soluble reductants such as glutathione, fat-soluble vitamins and enzymes such as glutathione peroxidase and superoxide dismutase. Under conditions of oxidative stress, inducibility of these enzymes is an important response to pollutant-induced stress (Sole, 2000). These nonspecific responses are dependent on the intensity and the duration of oxidative stress applied to the organisms. Thus, antioxidant parameters may be useful as biomarkers of complex contaminants in marine waters (Doyotte et al., 1997). Amongst the antioxidant responses, hepatic GSH and CAT were induced in most of the PAHs and OCs exposure regimes in this study (Table 2). During the entire exposure period, the induction of GSH was observed in four sampling weeks in regime B. GSH is an oxyradical scavenger. It also acts as a reactant in conjugation with electrophilic substances. Thus, changes in GSH levels are useful indicators of the detoxification ability of organisms. Increases in levels of hepatic glutathione have been reported in mussels exposed to PAHs in a laboratory study (Cheung et al., 2001); in catfish exposed to sediments contaminated by aromatic hydrocarbons (Di Giulio et al., 1993); and in flatfish collected from PAH and PCB-contaminated sites (Stein et al., 1992).

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Previous studies have found similar results for CAT when mussels were exposed to xenobiotics in laboratory and field experiments. Krishnakumar et al. (1997) found that CAT could be used as an indicator of PCB exposure in the mussel Mytilus edulis. Furthermore, a significant increase in CAT activity was observed in the digestive gland of ribbed mussels exposed to paraquat (Wenning et al., 1988); in M. edulis exposed to menadione (Livingstone et al., 1990); and in M. edulis exposed to PAH- and PCB-contaminated sediments (Porte et al., 1991). As stated by Cossu et al. (1997), catalase is often one of the earliest antioxidant enzymes to be induced. A similar result was observed in the present study, where CAT was induced in the first week when the target contaminants were added in exposure regimes A, B and C (Table 2). CAT may be considered as an important early indicator for oxidative stress, but may only be useful when sampling occurs during the immediate contaminant exposure period. GST activity in mussel gills exposed to a constant exposure (regime A) did not increase until the last week of exposure, whereas hepatic GST was induced in regime B in the last two weeks of exposure (Table 2). In the latter dosing regime, it is interesting that the induction of GST was observed during the depuration period. Another study (Moore et al., 1987) has described similar results in which the induction of mollusc (Mytilus sp.) enzymatic systems was observed during the depuration period. In the present study induction of GST activity occurred infrequently, and it may be that GST does not immediately respond to oxidative stress from short term exposure to these PAH and OC contaminants. There are mixed conclusions about the usefulness of GST as a biomarker of contaminant induced stress in the literature. Significant differences in GST activity were noted in M. edulis and Littorina littorea from different exposure systems (Lee, 1988; Suteau et al., 1988). However, inconsistent results were observed in fish when exposed to different pollutants (Van Veld et al., 1988; George, 1989; Collier and Varanasi, 1991; Di Giulio et al., 1993). In the present study, it may be that toxicant concentrations in some exposure treatments were high enough to inhibit GST activity or perhaps some toxic intermediates produced during xenobiotic metabolism may have inactivated GST and thus resulted in inconsistent GST activity in mussels. Several studies have detected enhanced lipid peroxidation in aquatic organisms exposed to high concentrations of xenobiotics in water (Gabryelack and Klekot, 1985; Viarengo et al., 1988, 1989; Wenning and Di Giulio, 1988; Wenning et al., 1988; Bano and Hasan, 1989; Ribera et al., 1991; Thomas and Wofford, 1993) and in sediments (Di Giulio et al., 1993; Sole et al., 1996). In the present study (see Table 2), the induction of lipid peroxidation was observed in gill, but not in hepatopancreas, when mussels

were exposed under constant dosage (regime A). The extent to which oxyradical generation produces biological damage is dependent on the effectiveness of antioxidant defences (Diguiseppi and Fridovich, 1984; Michiels and Remacle, 1988); i.e., lipid peroxidation is likely to be observed in the absence of sufficient antioxidant defence (Kappus, 1985). In the present study, higher antioxidant defences were observed in hepatopancreas than in gill. Therefore, it is unsurprising that induction of lipid peroxidation was observed in gill tissue. This result supports the argument that high levels of lipid peroxidation can be related to low activity of antioxidant defences systems. In ‘‘Musselwatch” programs, concentrations of chemical contaminants in the mussel tissues are used as an indication of the biological availability of chemicals in the water at a particular time and place or to predict long-term temporal trends in water quality (Goldberg, 1986). The bioaccumulation patterns are related to the route of uptake of xenobiotics, time course of exposure and physiology (Sole et al., 1994). Mussels exhibit a preferential uptake by direct partitioning from water (Porte and Albaiges, 1993). Therefore, changes in concentrations of contaminants in bivalve tissue can provide a time-integrated measurement of the contaminant bioavailability in the water. Furthermore, body burden data can assist with interpretation of biomarker data. Body burdens of PAHs in the constant exposure regime (regime A; Fig. 1a) showed an excellent correlation with their Kow valves. However, such a correlation was not observed in other dosing regimes. Several studies have reported that the tendency for accumulation of organic contaminants in mussels can be correlated with Kow values of compounds only in a simple water system (Geyer et al., 1982; Mackay, 1982), and in a sediment-water system in which the direct source of contaminants is the dissolved phase (Pruell et al., 1987). According to Richardson et al. (2005) and Baumard et al. (1999), PAHs can be distributed 98% in particulate phase while dissolved PAH concentrations are often very low (100–200 ng/L). Neff and Burns (1996) stated that PAHs with high Kow require a long time to reach equilibrium in tissues of aquatic animals. The use of a Kow approach has been used as a common model for evaluating contaminant uptake in biota (Isnard and Lambert, 1988; MacKay et al., 1980). However, Kow is not necessarily the sole factor that affects contaminant uptake in aquatic animals. Several studies have reported that this approach may not provide accurate predictions for all types of chemicals (Gobas et al., 1989), particularly for the extremely hydrophobic chemicals. Different feeding rates, rapid elimination of chemicals into faeces, metabolic transformation within the body and adsorption onto organic matter in water may be some possible causes in reduced bioavailable fractions of

Table 4 Correlation coefficients (r-values)a for significant correlations between antioxidative responses in hepatopancreas and target pollutants with all mussel samples Compound

GPx

GR

SOD

GST

GSH

LPO

CAT

Anthracene Fluoranthene Pyrene B[a]P Total PAHs HCH Aldrin Dieldrin p,p0 -DDT Total OCs

ns ns ns ns ns ns 0.173* ns ns ns

0.166* ns ns ns ns 0.207* 0.203* 0.292*** 0.353*** 0.321***

0.209* 0.179* 0.230** 0.146* 0.247** 0.167* ns ns 0.203* ns

ns ns ns 0.647* ns ns 0.250* ns 0.558* ns

0.270** 0.199* 0.254** 0.243** 0.300*** 0.150* 0.289*** 0.142* 0.219** 0.171***

ns ns ns ns ns ns ns ns ns ns

0.245*** ns 0.234*** ns 0.244*** 0.493*** 0.317*** 0.227** 0.214* 0.214*

ns: P > 0.05. a Non-parametric Spearman correlation coefficient. * P < 0.05. ** P < 0.01. *** P < 0.001.

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Table 5 Correlation coefficients (r-values)a for significant correlations between antioxidative responses in gill and target pollutants with all mussel samples Compound

GPx

GR

SOD

GST

GSH

LPO

CAT

Anthracene Fluoranthene Pyrene B(a)P Total PAHs HCH Aldrin Dieldrin p,p’-DDT

0.208** 0.360*** 0.289*** 0.198** 0.307*** 0.115* 0.065* 0.116* 0.164*

ns ns ns 0.143* ns 0.130* ns ns ns

ns ns ns ns ns ns ns ns ns

ns ns ns 0.244** ns ns ns ns ns

ns ns 0.164* ns ns 0.201* 0.151* 0.239** 0.248**

ns ns 0.117* 0.177* ns 0.140* 0.198**

Total OCs

0.155*

ns ns ns ns ns ns ns 0.226** 0.183* 0.209**

ns

ns

ns

0.261**

ns

a

ns ns

Non-parametric Spearman Correlation Coefficient; *P < 0.05, **P < 0.01, ***P < 0.001 ; ns P > 0.05.

chemicals, especially in closed or experimental systems (McCarthy, 1983; Bruggeman et al., 1984; Landrum et al., 1985; Gobas et al., 1988). In regimes B, C and D, accumulated levels of anthracene, fluroanthene and pyrene, but not B[a]P, were well correlated with Kow values (Fig. 1b–d). Van Hattun et al. (1998) indicated that lipid based BCFs increased with increasing hydrophobicity (Kow). However, these authors observed that for PAHs with log Kow greater than 6 (e.g. B[a]P; see Table 3), BCFs were lower than the expected values. This finding was confirmed in the present study. Measurement of the feeding rate of mussels could improve the understanding of the chemical uptake by organisms. Furthermore, a nonlinear correlation approach to evaluate the broader ranges of hydrophobicity could be derived to more accurately predict contaminant uptake with chemicals having a log Kow greater than 6. Concentrations of OC pesticides in mussels declined during the last two weeks of exposure in regime A (Fig. 1a). As revealed by Ding and Wu (1995), p,p0 -DDT may be partially transformed to p,p0 -DDD and p,p0 -DDE. This might explain the reduction of tissue p,p0 -DDT concentrations in mussels exposed to regime A. One assumption in the present study was the delivery of the same total amount of contaminants in each regime after 28 days exposure (see Table 1). However, significant differences in body burden of total PAHs and OCs under similar regimes were observed (i.e. regimes B and C; Fig 1b and c). This was attributed to the considerable loss of pollutants during the ‘‘no dose” period in the last 2 weeks of regime B. Loss of tissue PAHs during depuration periods was observed in dosing regimes B and D. This is possibly due to the loss of unassimilated food from the digestive gland and subsequent diffusion out of mussel tissue (Richardson et al., 2005). However, only a slight decline of OC pesticide concentrations in mussel tissue during the depuration weeks was observed in regime D (Fig. 1d). This may be due to the strong partitioning abilities of OC pesticides in the lipids of living organisms. Furthermore, different depuration periods may affect the body contaminant levels in mussel tissues. With reference to the present results, significantly different total tissue OC concentrations were observed between regimes B and D (Fig. 2). This result can be explained by different depuration periods in regime B (14 days) and D (7 days). The longer the depuration period, the lower the resulting body burden. Therefore, the mean OC concentration in regime B (37 ng g1 lipid weight) was lower than that in regime D (100 ng g1 lipid weight). According to Kahn (1977), about 50– 90% of absorbed chlorinated pesticides can be depurated within 4 weeks. As indicated by Richardson et al. (2001, 2005), PAHs and OC pesticides in mussel tissue (except p,p0 -DDT) can be depurated by over 95% within 10 days. Similar results were revealed by Wang (1998) who showed a positive relationship between PCB levels in fish and exposure duration. The aforementioned results showed

that exposure duration and dosage levels are among the key factors affecting body burdens of contaminants in mussels. Cheung et al. (2001) demonstrated that GSH in the mussel P. viridis may be the first antioxidative response. In the present study, GSH levels increased in mussels when the tissue PAHs or OCs concentration was below 60 ng/g lipid weight. This may be in order to balance pro-oxidant forces when mussels are exposed to electrophilic xenobiotics. However, when tissue OCs concentration was over 60 ng/g lipid weight, a decrease in GSH level was observed. Another finding was that gill GSH levels were negatively correlated with total tissue OC pesticides after 28 days of exposure (Table 5). This may be due to the formation and excretion of GSH-conjugates resulting in a net loss of GSH from cells (Hageman et al., 1992). In contrast, hepatic GSH showed positive relationships with all target pollutants (Table 4) while only B[a]P was similarly associated with gill GSH. Moreover, hepatic GSH exhibited positive correlations with total OCs and total target pollutants. In previous laboratory studies, a significant decrease in GST activity has been observed in cytosol and gills of mussels exposed to B[a]P (Michel et al., 1993; Akcha et al., 2000). In this study, a negative correlation between tissue B[a]P and GST activity was also observed in hepatopancreas. However, Fitzpatrick et al. (1995) reported a slight increase of GST activity with increasing tissue concentration of PAH in the digestive gland in molluscs. In other field and transplantation studies, no correlations between GST activity and PAH pollution levels were found in mussel digestive gland or gills (Livingstone et al., 1995; Sole et al., 1996; Fitzpatrick et al., 1997). Similar results were observed in the present study after 28 days exposure in which there was no correlation between tissue contaminant burden and gill/hepatic GST activity (Tables 4 and 5). These contradictory results may suggest that GST is not an effective biomarker of the PAH or OC pesticides evaluated in the current study. GPx is an important enzyme involved in removing hydrogen peroxide and lipid hydroperoxides and its role has been investigated in different types of marine invertebrates such as molluscs (Gamble et al., 1995). Positive correlations have been observed between GPx activity and organochlorine body burden (i.e. PCBs, DDT and lindane) in mussels (Sole et al., 1994). In the present study, however, negative correlations between gill GPx and contaminants were observed (Table 5). Similar relationships were also found between gill GPx activity and each of the two target contaminant groups (total PAH and OC pesticides), but no significant correlations were observed for hepatic GPx activity and target contaminants, with the exception of Aldrin (Table 4). A decrease in GPx activity following organochlorine exposure has also been demonstrated by Videla et al. (1990). This may be due to the inhibition of enzyme synthesis by pesticides (Bainy et al., 1996). A negative correlation may also be a result of enzyme inactivation caused by high tissue contaminant concentrations (Borg and Schaich, 1984).

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According to Kappus (1985), both GPx and CAT catalyse transformation of hydroperoxide to molecular water; i.e., they may act on common substrates. Thus, there is competition for the same group of substrates. This may explain the low GPx activity observed in the present study. In the present study, positive correlations between CAT activity and tissue contaminant concentrations were demonstrated (Tables 4 and 5). Similar correlations have previously been found between CAT activity and tissue PAH concentrations in mussel hepatopancreas in field and laboratory studies (Akcha et al., 2000; Sole et al., 1998; Porte et al., 1991). Increased hepatic CAT activity was also observed in Scorpionfish exposed to PCBs (Rudneva-Titova and Zherko, 1994); in mussels exposed to a mixture of PAHs and PCBs (Krishnakumar et al., 1997) and in fish exposed in the field (Regoli et al., 1998). These results show that CAT activity could be a useful biomarker in both laboratory and field studies. SOD activity was found to be elevated in liver and gill of carp exposed to paraquat (Vig and Nemcsok, 1989); in hepatopancreas of mussels exposed to PAHs (Sole et al., 1994); and in hepatic postmitochondrial supernatant fractions in Spot (Leiostomus xanthurus) exposed to a PAH-contaminated environment (Roberts et al., 1987). In this study, hepatic SOD activity was found to be positively correlated with individual tissue PAHs and total PAHs (Table 4). A positive correlation between hepatic SOD and CAT was observed in the present study, probably because of the relationship between their functions (Amstad et al., 1994). As stated by Kappus (1985), SOD catalyzes the dismutation of the superoxide radical into hydrogen peroxide which is then detoxified by catalase (Halliwell and Gutteridge, 1989). An increase of SOD activity should, therefore, be associated with an increase in H2O2 production resulting in elevated CAT activity. When antioxidant defenses are impaired, oxidative stress may produce lipid peroxidation in cell constituents (Halliwell and Gutteridge, 1989). Previous studies have shown that the level of lipid peroxidation increases in the digestive gland of M. edulis and L. littorea following exposure to phenanthrene (Moore, 1988); in the digestive gland of G. demissa exposed to paraquat (Wenning et al., 1988); and in the mantle microsomes of M. edulis (Gultekin et al., 2000). In the present study, lipid peroxidation demonstrated no relationships with target contaminants in hepatopancreas after 28 days of exposure. However, negative correlations were detected between each individual and total OC pesticides and gill lipid peroxidation for all samples regardless of regimes and exposure durations (Table 5). Induction of gill lipid peroxidation was observed in regime A only (Table 2), which was probably due to comparatively low gill antioxidant enzyme activity. The level of lipid peroxidation is affected not only by the tissue contaminant levels, but also by the antioxidant ability of an organism. Therefore, the level of antioxidant force and tissue contaminants must be taken into account during data interpretation. Exposure of channel catfish to PAH-contaminated sediment for up to 28 days resulted in consistent increases in hepatic CAT activity, GSH and GSSG levels, and lipid peroxidation levels (Di Guilio et al., 1989). Ribbed mussels exposed to paraquat for up to 96 hours exhibited significant increases in SOD, CAT and GSH level (Wenning et al., 1988). In the present study, a generally good relationship was observed between antioxidant responses and tissue OC pesticides. This may be due to the persistence of OC pesticide levels in the lipid of mussel tissues. However, such a relationship did not exist for the PAHs, which are comparatively more easily depurated and biotransformed than OC pesticides. Hepatic GSH and CAT in regimes A and B showed a very clear positive correlation with tissue OC pesticides (Table 4). In this study, most of the antioxidant parameters were induced in mussels exposed to regimes A and B (Table 2), but not regimes C and D in

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which mussels had a higher body burden of contaminants (Fig. 2). Stegeman et al. (1992) reported that enzyme activity might be inhibited once contaminant burdens reach a certain level in body tissue. This is a plausible explanation for the lower levels of antioxidant response observed in regimes C and D. The overall objectives of the present study were to examine the behaviour of commonly used antioxidant enzymes in green-lipped mussels (P. viridis) exposed to various exposure regimes, and to determine the correlation between enzyme activities and the toxicants. This study demonstrated that hepatic GSH and CAT are least affected by different dosing patterns (especially for OC contamininats), as they generally exhibited a similar induction response over various exposure regimes. This sets an important criterion for selecting an appropriate antioxidant enzyme in future studies, although if GSH and CAT are to be considered as reliable biomarkers, and not affected by exposure regime, one would expect at the end of the exposure period (i.e., 28 days), they would be significantly induced in the different regimes as all mussels at that sampling point were exposed to the same amount of PAHs and OCs. Contaminant uptake by mussels exposed to different regimes did not completely follow the commonly used Kow approach. Although results of the present study revealed that bioaccumulation of target PAHs generally followed the pattern predicted by Kow estimation, an exception was observed for B[a]P, where levels were lower than the expected values. For OCs, the problems of biotransformation should be considered when interpreting the results, especially for p,p0 -DDT which may be partly transformed to p,p’-DDE and p,p’-DDD. The results also revealed that pollutants behaved differently during the depuration periods. Exposure duration and dosage levels are probably the key factors affecting pollutant depuration rates, as well as burden of contaminants. Furthermore, the difference in lipophilicity of each toxicant may affect the depuration rates. In the present study, it was interesting to note the important relationships between antioxidant enzymes, such as the positive correlations between hepatic SOD and CAT, and hepatic GSH and GPx after 28 days exposure. These relationships are probably the result of interactions within the antioxidant defence system. A complex system of antioxidant defence has evolved to protect organisms against oxidative stress, and it is recommended that further studies be undertaken to investigate the inter-relationships between different antioxidant enzymes. Results of this experiment (e.g., see Table 2) indicated that most of the antioxidant enzymes were induced under exposure to regime A (constant dose) and regime B (initial pulse followed by no exposure). Hepatic CAT and GSH in regimes A and B demonstrated positive correlations with tissue total OC pesticides (Tables 4 and 5). These findings provide direction and criteria for further studies. The continuously dosed regimes produced larger oxidative stress than pulsed regimes. However, it must be concluded that antioxidant parameters may not be useful indicators of contaminant exposure under conditions where ‘‘no dose” periods are a component of the experimental design or are an environmental fact. This has considerable implications for field studies, particularly under conditions where ambient contaminant concentrations may vary considerably. Our results suggest that the use of antioxidant biomarkers under these circumstances is, at best, of limited value for environmental impact or risk assessments. In conclusion, hepatic GSH and hepatic CAT were observed to be potentially effective biomarkers of exposure to mixtures of organic pollutants, showing correlations with tissue PAHs and OC pesticides in P. viridis under certain dosing regimes, and at various stages of the experiment. However, before they can be routinely used as monitoring tools, additional laboratory and (perhaps especially) field studies need to be undertaken over selected (i.e., longer vs. shorter) exposure periods.

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