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Relationship between biomarker responses and contaminant concentration in selected tissues of flounder (Platichthys flesus)fromthe Polish coastal area of the Baltic Sea
Baltic Sea Contaminants Flounder Biomarkers
Dorota Napierska Magdalena Podolska*
Sea Fisheries Institute,
Kollataja 1, PL-80-332 Gdynia, Poland;
* corresponding author
Received 7 February 2008, revised 26 June 2008, accepted 23 July 2008.
Previous studies in the Gulf of Gdansk discussed the responses of selected enzymatic biomarkers to the contaminant gradient in fish and mussels. In the present study, flounder muscle and liver tissues were analyzed for polychlorinated biphenyls (PCB congeners: 28, 52, 101, 118, 138, 153 and 180), organochlorine pesticides (HCHs, HCB and DDTs), and trace metals (Pb, Cd, Zn, Cu, Hg, Cr). An attempt was made to identify the relationship between the measured enzymatic biomarker responses (cholinesterases, malic enzyme, isocitrate dehydrogenase, glutathione S-transferase) and contaminant concentrations in selected flounder tissues. The observed differences in enzymatic biomarker levels suggest that chronic exposure to low-concentration mixtures of contaminants may be occurring in the studied area. However, no conclusive evidence was found of a clear link between the biomarker responses and contaminant concentrations in flounder tissues.
Over the past fifty years there have been substantial inputs of POPs into the Baltic Sea from numerous sources. These include industrial discharges of organochlorines in effluents from pulp and paper mills, runoff from farmland, the antifouling paints used on ships and boats, and dumped
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wastes. Although recent monitoring data indicate that loads of certain hazardous substances have been reduced over the past ten years, problems do persist (HELCOM 2002). Concentrations of dioxins and PCBs in marine ecosystems declined in the 1980s, but this decrease leveled off in the 1990s. The concentration of dissolved trace metals in the Baltic Sea is many times higher than in the northern Atlantic, and these elements can still be found in high concentrations in certain marine organisms (HELCOM 2003). For endocrine disrupting substances and new contaminants like flame retardants, a full assessment of their levels or effects is not possible owing to the lack of monitoring data.
It has been suggested that the biological effects of such contamination may be more severe in brackish water systems than in marine systems. Laboratory studies have indicated that the sensitivity of aquatic animals to contaminants is closely related to the disruption of water and ion regulation in them (Cleveland et al. 1991, Dave et al. 1993). The results of the EU BEEP Project clearly indicated that the present contaminant concentrations in the different parts of the Baltic Sea are eliciting biological responses in various species, and in some cases are leading to chronic stress (Barsiene et al. 2006, Lang et al. 2006, Schiedek et al. 2006, Vuorinen et al. 2006). The Gulf of Gdansk is one of the anthropogenically most severely stressed sites in Polish marine areas because of the great impact of the Vistula River, the second largest river in the Baltic Sea drainage area. Field-testing of a battery of biomarkers in this location showed responses of selected enzymatic biomarkers to a contaminant gradient in both fish and bivalves (Napierska & Podolska 2005, Kopecka et al. 2006, Napierska et al. 2006).
Many substances can affect the physiological processes of living organisms through the induction or suppression of enzymatic reactions (Stegeman et al. 1988, Galgani et al. 1992, George 1994, Escartin & Porte 1996, Kirby et al. 2000). The neurotoxic effects of carbamates or organophosphates can be evaluated by measuring the inhibition of cholinesterase (ChE) activity; however, many authors have also linked the inhibition of ChE activity to the toxic effects of heavy metal exposure (Payne et al. 1996, Bocqueneet al. 1997, Guilhermino et al. 1998, Dethloff et al. 1999, Sturm et al. 1999). Widespread pollutants such as PAHs, PCBs and dioxin-like compounds usually induce phase II biotransformation glutathione S-transferase (GST) activity (Gubbins et al. 2000, Lenartova et al. 2000, Porte et al. 2000, Xu et al. 2001). Far fewer studies have described changes in enzyme activity that reflect basic metabolism in relation to exposure to contaminants. The malic enzyme (ME) that occurs in the cells of the metabolically more active tissues in crustaceans and bivalves may play an important role in pyruvate and Krebs cycle intermediate metabolism (Paynter et al. 1985,
Brodey & Bishop 1992a,b). It has been suggested that extramitochondrial malic enzyme could be one of the enzymes involved in the anaplerotic supply of Krebs cycle intermediates in skeletal muscle (Swierczynski 1980, Biegniewska & Skorkowski 1983, Konradt & Braunbeck 2001). Isocitrate dehydrogenase (IDH) is an enzyme that participates in the citric acid cycle. Both NADP-dependent enzymes have the ability to regenerate cellular NADPH, which is a necessary cofactor in antioxidant and detoxification systems.
In earlier publications, the authors presented the results of enzyme biomarker measurements (ChEs, GST and EROD) in flounder from the Polish coastal area of the Baltic Sea (Napierska & Podolska 2005, Napierska et al. 2006). In the present study, the results of previous measurements of ChE and GST activity and new data on contaminant concentration in selected tissues of flounder are analyzed using the new GLM model. The activities of two other enzymes (ME, IDH) are also discussed. An attempt is made to establish a relationship between all the measured biomarker responses and contaminant concentrations in selected flounder tissues.
2. Material and methods
2.1. Sample collection and handling
Flounder were sampled in the southern Baltic Sea in September 2001, 2002 and 2003. The fish were caught at three sites regarded as 'contaminated' in the Gulf of Gdansk (C1 - Mechelinki, C2 - Sopot, C3 - the Vistula Mouth) and from a site on the open sea coast, considered to be 'clean' and designated as the reference region (REF - Leba) (Figure 1). Detailed descriptions of the study area in relation to contamination can be found in the authors' earlier publications (Napierska & Podolska 2005, Napierska et al. 2006). Thirty fish were collected at each site (15 males and 15 females) during each sampling; only fish over 20 cm in length were taken. The liver and muscle of each fish were excised and immediately frozen at — 80°C for biochemical analysis. Dissections were performed within 1 h of capture. The total body length (cm) weight (g), age, sex and gonad developmental stage of each fish were recorded. The age was determined from otoliths, and the gonad stages were classified according to Maier's scale. Fulton's formula was used to determine the body condition factor, CF:
CF = J x 100, where w - total weight and l - length of the fish.
Somatic indices for liver (HSI) and gonads (GSI) were determined as follows: HSI = — x 100
GSI = ^ x 100, Wf
where wi - weight of the liver, wg - weight of gonads and wf - weight of fish.
17o00' 17o20' 17o40' 18o00' 18o20' 18o40' 19o00'
Figure 1. Location of flounder (Platichthys flesus) sampling sites: REF - 'clean' reference site, C1, C2, C3-contaminated sites (Cl-Mechelinki, C2-Sopot, C3 - Vistula Mouth)
2.2. Preparation of tissue homogenates and enzyme activity determination
The methods of extraction of ChEs and GST were described earlier (Napierska & Podolska 2005). ChE activities were determined using the method described by Ellman et al. (1961) and adapted for use with a microplate reader (Bocqueme & Galgani 1998). The enzyme kinetics was monitored at 412 nm for three minutes. The standard reaction mixture (final volume 0.380 cm3) contained 0.02 M phosphate buffer (pH 7.0), 0.5 mM DTNB (5,5'-dithiobis(2-nitrobenzoic acid)) and 2.6 mM ACTC (acetylthiocholine chloride) or BCTC (butyrylthiocholine chloride).
GST measurements were performed using a modification of the method described in Habig et al. (1974). The enzyme activity was tracked
spectrophotometrically with a microplate reader at 340 nm. The standard reaction mixture (final volume 0.210 cm3) contained 0.1 M phosphate buffer
(pH 7.4), 1 mM CDNB and 1 mM GSH.
For IDH and ME, extraction was performed on 500 mg muscle tissue using a 0.02 M phosphate buffer (pH 7.0). The tissue was homogenized in buffer (1:4, w:v) and centrifuged at 10 000 g for 20 minutes at 4°C. An aliquot of the supernatant (the S9 fraction) was stored at —80°C and used in the assay. IDH and ME were assayed spectrophotometrically as described in Gronczewska et al. (2003) by monitoring the changes in absorbance at 340 nm as a result of the appearance of NADPH. The standard reaction mixture for the ME assay contained 50 mM Tris-HCl at pH 7.5, 0.5 mM NADP, 10 mM L-malate and 1 mM MnSO4. The standard reaction mixture
for the IDH assay contained 50 mM Tris-HCl at pH 7.5, 0.5 mM NADP,
5 mM isocitrate and 1 mM MnSO4.
Protein concentration was determined as described by Bradford (1976) using the Protein Kit II from Bio-Rad laboratories and bovine serum albumin as the protein standard.
2.3. Chemical analysis
The flounder muscle tissue from the females from one sampling site was pooled and analyzed for polychlorinated biphenyls (PCB congeners:
28, 52, 101, 118, 138, 153 and 180), organochlorine pesticides (HCHs,
HCB and DDTs) and trace metals (Pb, Cd, Zn, Cu, Hg, Cr). The same pooling procedure was applied to the muscle tissue from the males, and to the male and female liver tissue. Liver and muscle tissue was lyophilized and extracted with hexane to obtain a fat containing the accumulated organochlorine compounds. This extract was cleaned with sulfuric acid (di Muccio et al. 1990). DDTs and PCBs were determined by capillary gas chromatography with electron capture detection (UNEP/IOC/IAEA 1988). Trace metals were determined by atomic absorption. The fish tissue was mineralized with nitric acid in microwave ovens. The zinc and copper concentrations (> 1mgkg_1) were determined by the flame method, and the concentrations of lead, cadmium, chromium and copper (< 1mgkg_1) were determined by the flameless method in a graphite furnace. The mercury concentration was measured by the vapor generation method in a gold amalgam mercury analyzer (AMA 254) (UNEP/FAO/IAEA/IOC 1984a,b). The Testing Laboratory of the Sea Fisheries Institute works in compliance with a quality system (the laboratorys competence is confirmed by Accreditation Certificate No. AB 017).
2.4. Statistical analysis
The significance of the differences between average concentrations of trace metals and organochlorine compounds in the muscle tissue and liver of flounder sampled at the clean and contaminated sites were tested with the non-parametric Mann-Whitney U-test. The significance level was taken to be p < 0.05. Next, the data were analyzed using generalized linear models (GLM) (McCullagh & Nelder 1989). The calculations were performed using the GenStat software package (GenStat 2002). Enzymatic activity (AChE, BChE, GST, IDH, ME) was modeled as being dependent on the area, year of sampling, the biological parameters of the fish (body length, condition factor, sex, gonad maturity stage) and contaminant concentrations as the explanatory variables. The values of the dependent variables were log-transformed. Separate models were fitted for each of them:
y = area + year + sex + gonad stage + LT + CF + Cd + Cr + Cu + + Hg + Pb + Zn + DDTs + HCHs + PCBs + HCB + error
where y (dependent variable) - enzymatic activity in muscle tissue (AChE, BChE, IDH, ME) and liver (GST); LT - total length of fish; CF - body condition factor; Cd + Cr + Cu + Hg + Pb + Zn + DDTs + HCHs + PCBs + HCB - contaminant concentrations in fish tissues (muscle or liver). The concentrations of contaminants in fish muscles were used for modeling the muscular enzymatic activity (AChE, BChE, IDH, ME), and the contaminant levels in fish liver for hepatic GST. The area has 4 factor
levels (REF, C1, C2, C3), the year 3 levels (2001, 2002, 2003), the sex 2
levels (males, females) and the gonad stage 4 levels (2, 3, 4, 8).
The error term was assumed to be normally distributed with zero mean and constant variance. Corner point parameterization was used, i.e., factor effects for level one were assumed to be zero for all factors. Thus, the factor effect for the other levels can be regarded as the difference between the effect at a given level and the effect at level one. The significance of factors and variables was tested, and only significant terms were left in the final model. Similarly, factor levels that did not produce a significantly different response of the enzymatic activity were grouped into new factor levels. The tests were performed by deletion, i.e., only those terms whose deletion did not result in a significant increase in deviance (i.e., the GLM measure of discrepancy between modeled and observed values) were left in the model, and the F-test was used. The model assumptions and performance were evaluated by analyzing residuals.
3.1. Biological parameters
Table 1 shows the biological parameters of the examined fish. Fish length ranged from 23to 41 cm. The fish sampled at Leba (the 'clean', reference site) were usually longer, with a higher condition factor in comparison to the fish sampled at the other sites. The HSI and GSI index values were usually also the highest in fish from this area.
Table 1. Biological parameters recorded in the examined flounder (Platichthys flesus)