- Email: [email protected]
Organochlorine bioaccumulation and biomarkers levels in culture and wild white seabream (Diplodus sargus)
Chemosphere 73 (2008) 1669–1674
Contents lists available at ScienceDirect
Chemosphere j o u r n a l h o m e p a g e : w w w . e l s e v i e r. c o m / l o c a t e / ch e m o s p h e r e
Organochlorine bioaccumulation and biomarkers levels in culture and wild white seabream (Diplodus sargus) Marta Ferreira a,*, Paulo Antunes b,d, Joana Costa a, Joana Amado b, Odete Gil b, Pedro Pousão-Ferreira c, Carlos Vale b, Maria Armanda Reis-Henriques a,d a CIMAR/CIIMAR – Centro Interdisciplinar de Investigação Marinha e Ambiental, Universidade do Porto, Laboratório de Toxicologia Ambiental, Rua dos Bragas, 289, 4050-123 Porto, Portugal b INRB/IPIMAR – Instituto Nacional dos Recursos Biológicos, IPIMAR, Av. Brasília, 1449-006 Lisboa, Portugal c INRB/IPIMAR SUL – Instituto Nacional dos Recursos Biológicos, Av. 5 de Outubro, 8700-305 Olhão, Portugal d ICBAS/UP – Instituto de Ciências Biomédicas Abel Salaz ar, Universidade do Porto, Largo Professor Abel Salazar, 2, 4099-003 Porto, Portugal
a r t i c l e
i n f o
Article history: Received 24 March 2008 Received in revised form 13 June 2008 Accepted 24 July 2008 Available online 11 September 2008 Keywords: PCB DDT Fish farming EROD GST Micronucleous test
a b s t r a c t Persistent organic pollutants (POPs), which can accumulate in the adipose fish tissues, can enter the human food chain through the consumption of fish, and cause risk to health. The use of chemical analysis, and biochemical and cellular responses is a way to detect the impact of pollutants in aquatic systems. The purpose of this study was to investigate the levels of organochlorine compounds (polychlorinated biphe nyls – PCB and p,p9-dichlorodiphenyltrichloroethane and its metabol ites – tDDT) in, wild and cultivated, white seabream (Diplodus sargus), and also its biological effects that were evalua ted by assessing the activ ity of biotransformation enzymes and genotoxic effects. To achieve that we have sampled five different size classes (I – 13 g, II – 64 g, III – 143 g, IV – 315 g and V – 441 g) of white seabream from a local aquacul ture, and also a group of wild fish (375 g) in order to compare accumulation and responses between cul tured and wild fish. White seabream, cultured and wild, presented low levels of organochlorine content, both in liver and in muscle. Wild white seabream, in comparison to cultured ones at the marketable size, showed lower organochlorine accumulation. Biotransformation enzymes showed negative correlations with organochlorine levels in liver. Micronucleous numbers revealed that wild white seabream are not so exposed to genotoxic compounds as cultured ones. © 2008 Elsevier Ltd. All rights reserved.
1. Introduction In recent years, there has been an increasing awareness of the need to assess the adverse effects of contaminants in aquatic organisms, and also the risk of fish consumption for human health, mainly regarding farmed fish. Persistent organic pollutants (POPs), such as organochlorines, are lipophilic compounds that bioaccu mulate and biomagnify through the food chain (Singh and Singh, 2008a,b). The most widespread organochlor ines in the environ ment and in animal tissues are polychlorinated biphenyls (PCB), dichlorodiphenyltrichloroethane (DDT), and specially the DDT deg radation product, dichlorodiphenyldichloroethylene (DDE) (Toft et al., 2003), and their levels found in some marine organisms gave rise to some concern. In the marine environment, these hydropho bic compounds accumulate in the adipose tissues of fish thus enter ing the food chain. More recently, some concern has arise on halogenated com pounds in aquaculture systems namely the potential hazards * Corresponding author. Tel.: +351 22 340 18 00; fax: +351 22 339 06 08. E-mail address: mferreira@ciimar.up.pt (M. Ferreira). 0045-6535/$ - see front matter © 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2008.07.070
associated with the ingredients used in aquaculture feeds (Easton et al., 2002; Jacobs et al., 2002; Antunes and Gil, 2004; Maule et al., 2007). Moreover, several studies have reported higher orga nochlorine levels in cultivated species in comparison with wild species, like salmon (Easton et al., 2002) and seabass (Antunes and Gil, 2004). Among the available techniques, the integrated use of chemical analysis and biochemical and cellular responses is a way to detect the impact of anthropogenic contaminants in aquatic systems, both wild and cultivated fish (Ferreira et al., 2004; Fernandes et al., 2007). Xenobiotic compounds may be biotransformed in liver by enzymes from phase I and phase II. Phase I is a non-synthetic alter ation (oxidation, reduction or hydrolysis) of the original foreign molecule, which can then be conjugated in phase II (Commandeur et al., 1995). Biotransformation of lipophilic compounds is a requirement for detoxification and excretion (Ferreira et al., 2006). However, certain biotransformation steps are responsible for the activation of foreign chemicals to reactive intermediates that ultimately result in toxicity, genotoxicity, or carcinogenicity (Van der Oost et al., 2003). Organochlorine compounds are potentially genotoxic, implying that they can damage DNA directly and/or
1670
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
enerate reactive species which can in turn damage DNA (Cachot g et al., 2006). The micronucleous (MN) test in fish has been shown to be a useful in vivo technique for genotoxicity testing and to have potential for in situ monitoring of water quality; the MN test detects micronuclei resulting from either chromosomal breakage during cell division or chromosome loss events in anaphase dam ages (Kim and Hyun, 2006). The marine aquaculture in the South-European countries is based on gilthead seabream (Sparus aurata) and seabass (Dicentra chus labrax) (Saavedra et al., 2006). However, the introduction of new species in aquaculture is an issue in development, and the white seabream (Diplod us sargus) is considered to be a promising new species with high market values and demand (Sa et al., 2006, 2007). The white seabream is an omnivorous species of the Sparidae family, found in the Mediterranean as well as in the eastern Atlantic Ocean, including the archipelagos of Madeira, Cape Verde and the Canary Islands (reviewed in Perez et al. (2007)). The purpose of this study was to investigate the levels of orga nochlorine compounds (tPCB and tDDT) in cultivated white sea bream (in different growth stages) and in the corresponding food pellets, and also the biological effects. The white seabream was chosen because of its high potential for aquaculture. This species is found in the Mediterranean, as well as in the Atlantic Ocean, and is mainly caught by self-employed fisherman and constitutes a valu able fishery resource considering its high price (Perez et al., 2007). The biological effects were evaluated by assessing the phase I and II biotransformation enzymes (EROD and GST, respectively) and genotoxic effects through the MN test. Identical study was also con ducted with the wild specim ens of white seabream. Moreover, to our knowledge the evaluation of the biotransformation enzymes activit ies, EROD and GST, widely used as biomarkers of exposure to pollutants, and the MN test to assess genotoxic effects, has never been reported in this particular species. 2. Materials and methods
weighted and the condition factor (CF) calculated (body weight (g) £ 100/length3 (cm)). Blood was collected from the caudal vein to the MN test. Liver and muscle were sampled for biochemical analysis and measurement of organochlorine compounds levels. The samples were transported to the laboratory in dry ice; samples for biochemical analysis were stored at ¡80 °C and for organochlo rine determinations at ¡20 °C. Water from the land tanks was filtered through pre-washed (hexane) and pre-combusted (350 °C, 17 h) Gelman A/E filters and particulate fraction collected. Filters were stored frozen and then dried at 40 °C for analysis. 2.2. PCB and DDT analysis 18 PCB congeners (IUPAC Nos. 18, 28, 52, 49, 44, 101, 151, 149, 118, 153, 105, 138, 187, 183, 128, 180, 170, 194), p,p9-DDT and metab olites (p,p9-DDD and p,p9-DDE) were quantified. Analyses were per formed according to Antunes and Gil (2004). Values are presented as the sum of the 18 congeners – tPCB, the sum of 7 indicator PCB namely PCB28, 52, 101, 118, 138, 153 and 180 – PCB7 (UNEP, 2007), and the sum of p,p9-DDT and its metabolites – tDDT. About 200 g of diet pellets was collected and homogenized. Subsamples of 2.00 g of diet pellets and muscle and 0.5000 g of liver were Soxhlet extracted with n-hexane for 6 h, and suspended particulate matter (SPM) samples were also Soxhlet extracted for 16 h. Lipid content was determined gravimetrically from aliquots of tissue extracts. The remaining extract was purified with a Florisil column and further with sulphuric acid before the analysis in an Agilent 6890N gas chromatograph, equipped with a micro elec tron capture detector and a DB-5 (J&W Scientific) capillary column (60 m £ 0.25 mm i.d. £ 0.25 lm film thickness). tPCB and tDDT were quantified using a six point calibration curve and CBs 65 and 204 as internal standards. Procedural blanks were analysed each 10–16 samples to monitor possible laboratory contamination. Detection limits calculated from three times the peak height in blank sam ples, ranged from 0.01 to 0.04 ng g ¡1 dw.
2.1. Sampling 2.3. Biochemical analysis White seabream samples were collected from a fish farm located in Ria Formosa, Olhão, in the South of Portugal, and wild specimens were captured in the coastal zone. The wild specimens sampled were correspondent to the higher weight classes, the marketable size, in order to compare the fish quality (wild and cul tured) for human consumption. Five different size classes (Table 1) according to their weight and a sub-sample of the food pellet correspondent to each size class were sampled. The two smaller classes were sampled from glass fibber tanks, with constant filtered water circulation, and the larger specimens from the land tanks with water circulation con trolled by the tidal cycle. The fish were killed by asphyxia in ice, and the tissues sampled at the fish farm laboratory. Wild animals were captured by anglers. The animals were immediately killed by asphyxia in ice and sampled in place. Animals were measured and
Table 1 Weight (g), length (cm) and condition factor (CF) for the different size classes of white seabream Class
N
Weight (g)
Length (cm)
CF
I II III IV V Wild
24 24 10 6 8 7
12.8 ± 0.3 63.7 ± 3.2 142.7 ± 5.8 315.5 ± 8.3 441.4 ± 11.9 375.3 ± 41.9
8.6 ± 0.1 14.8 ± 0.2 20.6 ± 0.4 25.2 ± 0.3 28.8 ± 0.3 27.6 ± 1.2
1.97 ± 0.03a 1.95 ± 0.03a,b 1.65 ± 0.05c 1.97 ± 0.10a,d 1.85 ± 0.03b,d,e 1.75 ± 0.05c,e
Values presented as mean ± SE. Different letters denote significant differences among the different groups, p < 0.05.
Livers were homogenized in ice-cold sodium phosphate buffer 50 mM, Na2EDTA 0.1 mM, pH 7.8 and centrifuged at 15000g for 20 min, at 4 °C. Glutathione S-transferase (GST) in the liver was determined according to the method of Habig et al. (1974) adapted to microplate as described in Ferreira et al. (2006) using glutathi one (GSH) 10 mM in phosphate buffer 0.1 M, pH 6.5, and 1-chloro2,4-dinitrobenzene (CDNB) 60 mM in ethanol prepared just before the assay. The reaction mixture consisted of phosphate buffer, GSH solution and CDNB solution in a proportion of 4.95 mL (phosphate buffer): 0.9 mL (GSH): 0.15 mL (CDNB). In the microplate, 0.2 mL of the reaction mixture was added to 0.1 mL of the sample, with final concentration 1 mM GSH and 1 mM CDNB in the assay. The GST activity was measured immediately every 20 s, at 340 nm, dur ing the first 5 min, and calculated in the period of linear change in absorbance. Liver GSH activity is expressed in nmol/min/mg protein. Liver ethoxyresorufin O-deethylase (EROD) activity was mea sured according to Ferreira et al. (2004). Briefly, liver was homog enized in ice-cold buffer (50 mM Tris–HCl, pH 7.4, 0.15 M KCl). Microsomes were obtained by centrifugation of the 9000g super natant at 36000g for 90 min in a SIGMA 3K30 centrifuge. The pellet was then resuspended in buffer (50 mM Tris–HCl, 1 mM NaEDTA, pH 7.4, 1 mM dithiothreitol, 20% v/v glycerol) and spun down at 36000g for 120 min (Fent and Bucheli, 1994). Microsomes were suspended in EDTA-free resuspension buffer and stored at ¡80 °C until use. Microsomal suspension (50 lL) was incubated with eth oxyresorufin 0.5 lM for 1 min, and the enzymatic reaction was
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
initiated by the addition of 45 lM NADPH. EROD activity was mea sured for 5 min at kex 530 nm and kem 585 nm, and determined by comparison to a resorufin standard curve. Hepatic EROD activity is expressed in pmol/min/mg protein. 2.4. Micronuclei test Two blood smear slides were prepared for each fish, fixated in methanol for 10 min and stained with Giemsa 5% in 3 mM phos phate buffer for 30 min. Two slides per fish were observed under a light microscope and micronuclei were recorded in a total of 1000 erythrocytes per slide. 2.5. Statistical analysis Differences between groups were tested using a One-Way ANOVA with a multiple comparison test (LSD) at a 5% significance level. Some data had to be log transformed in order to fit ANOVA assumptions. All tests were performed using the software Statis tica 6.0 (Statsoft, Inc., 2001). 3. Results The concentrations of organochlorine compounds in the food pellets and in suspended particulate matter (SPM) from the correspondent pounds are shown in Table 2. Values of the seven indicator PCB (PCB7) are also presented in Tables 2 and 3 to allow comparisons with other studies, the PCB7 represent about 65% of tPCB without significant differences to tPCB, therefore results will be discussed as tPCB. SPM showed low levels of tPCB and tDDT. The food pellets to feed animals from class III had the highest concentrations of these two types of organochlorine compounds. Table 2 Lipid content (%), sum of all analyzed PCB congeners – tPCB, sum of 7 indicator PCB congeners – PCB7, and sum of p,p9-DDT and its metabolites – tDDT, in the food pellets (ng g¡1 lipids) from each class of white seabream, and in the suspended par ticulate matter (SPM) (ng g¡1) Food pellet Classes I + II Class III Classes IV + V
Lipids (%) 19.4 16.9 8.9
tPCB ng g¡1 lipids 45 150 108
PCB7 ng g¡1 lipids 23 83 57
tDDT ng g¡1 lipids 26 45 21
SPM Class III Class IV
– –
ng g¡1 11 8
ng g¡1 5.4 3.7
ng g¡1 1.1 0.6
1671
The pellets of classes I + II presented the lowest levels of tPCB, how ever, the concentrations of tDDT were not different to the pellets supplied to fish from classes IV + V. Levels of lipid content, tPCB and tDDT in muscle and liver of cultivated and wild white seabream are displayed in Table 3. Ani mals from class I presented higher levels of lipid content in liver in comparison to the other classes although difference to class II and V were not significant. Lipid content in liver was not higher than in muscle, with the exception of individuals of class I. Regarding the organochlorine compounds, tPCB levels were always higher than tDDT. In white seabream from class II and wild animals tPCB levels were significantly higher in liver, in contrast with tDDT that were similar in both tissues. Muscle of individuals of class III showed higher levels of organochlorines than liver, and in the other size classes the two tissues showed no significant differences. Fish’s muscle from class III showed the higher content of tPCB and tDDT in comparison with the other analysed classes. Comparing wild animals with cultivated ones, of similar sizes (class IV and V), we can observe that wild seabream showed, in muscle and in liver, lower values of lipids, tPCB and tDDT than cul tivated animals (Table 3). In both tissues p,p9-DDT represented less than 27% of tDDT in farmed white seabream, and less than 35% in wild fish. Biomarkers evaluated in this study, hepatic EROD and GST activ ity, and MN in erythrocytes are presented in Table 4. White sea bream from class III showed significant higher levels of hepatic EROD activity, in comparison to the other classes analysed. This result is in agreement with the higher levels of organochlorine con taminants in the muscle. The fact that these fish had recently been transferred to the land tanks and been in the estuary waters could have influenced this biomarker. The lowest value for EROD activ ity was registered in animals from class V, which also showed the higher accumulation of tPCB and tDDT in liver. The wild specimens, that presented considerable lower levels of organochlorine accu mulation, showed EROD activity values similar to the ones from class II and IV. Overall, a significant negative correlation was found between tPCB and tDDT levels in liver with hepatic EROD activity. The accumulation of contaminants in liver and in muscle reflects different inputs; in muscle chronic accumulation is reflected while liver represents recent inputs. With this in mind we have correlated phase I enzyme with tPCB accumulation in muscle (Fig. 1). Inter estingly, a positive correlation was found for this enzyme activity and the tPCB accumulation in muscle; animals with lower weight showed higher levels of hepatic EROD activity and lower levels of tPCB in muscle; on the contrary, animals with higher weight reflected more tPCB accumulation and lower EROD activity.
Table 3 Lipid content (%), sum of all analyzed PCB congeners – tPCB, sum of 7 indicator PCB congeners – PCB7, and sum of p,p9-DDT and its metabolites – tDDT, in liver and muscle of white seabream in ng g¡1 lipids Class
n
I
6
II
6
III
10
IV
6
V
8
Wild
7
Lipids (%)
tPCB (ng g¡1 lipids)
PCB7 (ng g¡1 lipids)
tDDT (ng g¡1 lipids)
Liver
Muscle
Liver
Muscle
Liver
Muscle
Liver
Muscle
37.7 ± 6.7a (8.2–57.7) 27.1 ± 3.0a,b (18.3–38.8) 16.7 ± 1.9b,c (7.2–25.3) 14.3 ± 1.1c,d (9.7–17.4) 29.2 ± 3.3a (15.3–39.5) 9.7 ± 0.6d (8.1–12.2)
12.3 ± 0.9a,c (7.9–14.0) 20.5 ± 0.7b,d (18.4–20.9) 11.4 ± 1.6a,c (5.2–18.6) 19.3 ± 2.6a,b (7.6–25.8) 27.1 ± 2.3d (16.4–35.3) 8.5 ± 1.6c (2.9–13.9)
117.3 ± 4 4.5a (62.8–339.2) 212.6 ± 7.9a,b (185.5–242.2) 325.7 ± 56.6b (118.2–729.0) 297.1 ± 28.5b (191.4–380.8) 529.0 ± 42.5c (408.0–675.1) 156.8 ± 13.8a (116.7–215.6)
101.5 ± 4.8a (87.0–117.8) 167.2 ± 4.7b (153.2–183.1) 736.4 ± 66.4c (504.5–1016.3) 430.3 ± 70.3d (261.2–760.5) 471.4 ± 30.0d (358.5–649.0) 111.3 ± 12.6a (73.3–163.3)
78.3 ± 30.1a (40.3–228.0) 124.2 ± 3.6a,b (116.0–137.3) 222.9 ± 47.9c (81.5–482.5) 202.4 ± 20.1b,c (265.3–435.5) 338.3 ± 25.4d (265.3–435.5) 89.9 ± 7.9a (68.8–125.8)
57.9 ± 2.6a (51.3–66.9) 105.9 ± 3.4b (94.5–117.2) 500.0 ± 47.8c (315.9–698.7) 301.4 ± 49.8d (189.7–536.7) 323.8 ± 20.9d (243.0–443.7) 70.6 ± 9.1a (45.5–111.6)
59.3 ± 21.1a,b (30.4–164.1) 45.5 ± 2.8a,c (34.3–54.3) 42.2 ± 8.3a,c (7.7–86.0) 38.7 ± 6.3a,c (24.1–65.7) 77.8 ± 10.9b (44.4–134.7) 24.4 ± 2.0c (17.3–34.6)
34.0 ± 1.6a.b (28.1–39.1) 52.9 ± 2.0a,c (44.8–58.8) 103.2 ± 15.4d (52.8–66.4) 59.8 ± 9.9a,c (28.3–96.9) 71.7 ± 9.8c (45.6–124.5) 19.8 ± 9.8b (14.5–29.1)
Values presented as mean ± SE (minimum–maximum). Different letters denote significant differences among the different groups, p < 0.05.
1672
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
Table 4 Hepatic EROD and GST activities, and erythrocyte micronucleus (MN) in white sea bream Class
n
EROD (pmol/min/ mg protein)
GST (nmol/min/ mg protein)
MN (per 1000 erythrocytes)
I II III IV V Wild
6 6 10 6 8 7
80.0 ± 10.6a,c 48.3 ± 14.5a,b,d 125.1 ± 23.8c 44.0 ± 14.9b,d 21.6 ± 3.0d 55.1 ± 7.4a,b
104.2 ± 4.3a 108.4 ± 10.1a,b 118.8 ± 5.4a,b 127.2 ± 7.3b,c 76.8 ± 3.2d 148.8 ± 8.7c
2.1 ± 0.4a 2.0 ± 0.5a 2.5 ± 0.4a 1.0 ± 0.5b,c 1.9 ± 0.5a,b 0.3 ± 0.1c
Different letters denote significant differences among the different groups, p < 0.05.
Fig. 1. Correlation between hepatic EROD activity (pmol/min/mg protein) and tPCB (ng g¡1 lipids) in muscle in the lower classes (r) and in the higher weight classes (I).
Regarding hepatic GST activity the values are presented in Table 4, an increase in activity was observed from class I to IV. On the con trary, and in agreement with EROD activity, class V has also showed significant lower levels for this enzyme. The wild specimens pre sented significant higher levels for this biomarker. As for EROD activity a negative significant correlation was found between GST and tPCB in liver (r = 0.67; p < 0.05), and also with tDDT (r = 0.63; p < 0.05). The genotoxicity biomarker, evaluated as MN numbers, did not show significant differences between the evaluated classes, with the exception of class IV that showed lower values than the smaller classes of white seabream; nevertheless, class III has also presented a slightly higher value, as for EROD activity. Interestingly, and con trary to the other biomarkers assessed, wild specimens of white seabream presented significant lower numbers of MN, showing that wild animals are not so exposed to contaminants with geno toxic properties as animals from aquaculture. 4. Discussion The aim of the present study was to evaluate the levels of orga nochlorine contaminants (tPCB and tDDT) in a cultured species, the white seabream (D. sargus); and to detect and assess the possible adverse effects of these contaminants at a biochemical and cellular level, in different life stages of the species. In addition, to compare cultured specimens with wild specimens captured offshore. The observed differences between wild and cultivated fish may be explained by the different environments that result in different condition factors (Table 1). Cultivated fish have higher condition factor, as a result of the abundant food supply and closed environ ment. If we compare wild seabream only with the classes IV and V, both with a marketable size, the wild fish presented a lower
CF, the same has been observed for seabass wild and cultured, with wild seabass presenting significant lower CF (Fernandes et al., 2007). The fish farm is also located in a coastal lagoon, which contains higher levels of organic matter and contaminants than coastal waters (Quental et al., 2003). The higher levels found in farmed fish do not represent any risk to human consumption and are in the same order of other farmed and estuarine fish (Antunes and Gil, 2004). Excluding muscle of class III fish, tPCB levels presented an increase with fish size/age in both analysed tissues, and tDDT levels showed small variations. The fact that fish from class III presented lower lipid contents and higher organo chlorine levels than the other size classes is an unexplained result. One possible explanation is the higher content of organochlorine compounds in the food pellets supplied to the fish of this class (Table 2). However, the fact that this increase was not observed in liver make us consider other possibilities: (i) these fish were recently moved from indoor fibber glass tanks to outdoor land tanks, which could produce a stress factor that lead to an internal redistribution of contaminants in this adaptation period, this is in agreement with the lower condition factor found in this size class; (ii) there may be some differences in contaminant exposure due to sediment and suspended particulate matter; or (iii) there may occur variations in metabolic capacities of fish at this grow ing stage. Similar findings were obtained in another species, sea bass (D. labrax), produced in a similar fish farm (Antunes et al., 2007), with animals that still were in fibber glass tanks. This rein forces the idea of these variations being a normal biologic effect of fish growth. Indeed, these changes in metabolic capacities could explain higher hepatic EROD activity in white seabream class III, as described in Section 3. Biological effects in white seabream were evaluated by means of EROD and GST activity, and MN frequency. The referred biomark ers are not specific for PCB and DDT; however the integrated use of chemical and biological analysis is a way to detect the effects of exposure to contaminants (van der Oost et al., 2003; Ferreira et al., 2006). Phase I biotransformation enzyme, evaluated in this study by means of hepatic EROD activity has been described to be involved in the metabolic elimination of several contaminants, like PAH, PCB and others (Goksoyr and Forlin, 1992; Whyte et al., 2000; Ferreira et al., 2004, 2006). Liver is the organ where biotransforma tion of contaminants is processed; however we have found that ani mals presenting higher levels of OC accumulation, like the ones in class IV and V showed lower levels of EROD activity. In agreement with this are the wild animals that showed significant lower lev els of OC in liver and higher values for this enzyme. Same results were obtained in Atlantic tomcod from the Canadian coast, where the authors have found low hepatic EROD activity in correlation with higher levels of organochlorine contaminants (Couillard et al., 2005). Some other studies performed with largemouth bass, in a contaminated site with PCB has shown a moderate induction of EROD activity, and that after a short period of time enzyme levels have declined, suggesting a catalytic inhibition of this enzyme by certain PCB congeners (reviewed in Whyte et al. (2000)). Another possible explanation for the higher activity in smaller fish is that, usually smaller fish have a higher metabolic activity per gram (Whyte et al., 2000). Finally, we must not exclude the fact that hydroxylated PCB have been found in the environment and also in fish tissues (Buckman et al., 2006; Kunisue et al., 2007), indicat ing biotransformation from CYP enzyme-mediated, however the suggestion is that they could be biotransformed not by CYP1A but by CYP2B (Buckman et al., 2006), not measurable by hepatic EROD activity. One interesting results is the significant elevated EROD activity in class III. This could be a direct result of the transfer of the fish from the fiber glass tanks, with filtered water, to the land tanks with estuarine water that could have more EROD inducers present, like PAHs well documented EROD inducers (Van der Oost
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
et al., 2003; Ferreira et al., 2006). On the contrary, EROD activity showed a significant positive correlation with PCB accumulation in muscle (Fig. 1). Considering the muscle concentrations as a better indicator of total contamination in fish, adult white seabream pre sented a lower EROD activity response than juvenile. Taking into account only adults of commercial size, we obtained a positive cor relation between the hepatic EROD activity and PCB levels in the edible tissue, showing that EROD activity can be considered a good biomarker for total PCB contamination in fish. Glutathione S-transferase (GST) was assessed as phase II in the metabolic process of the biotransformation of contaminants. With exception of the fish from class V that presented significant lower levels of activity, the other weight classes presented sim ilar values between them. Overall, it was observed a significant negative correlation between this enzymes and the accumulation of OC. The decrease in this enzyme activity, in the presence of organochlorine contamin ants, has already been reported. A study performed with mullets has reported that before depuration, and with higher content in tPCB and tDDT, GST presented lower lev els of activity (Ferreira et al., 2006). Gallagher et al. (2001) have also stated that in brown bullhead, PCB have the ability to reduce GST expression. In fact the significant higher levels of GST activ ity presented by the wild white seabream than farmed seabass with simil ar weight (class IV and V), that showed considerable lower levels of OC accumulation can suggest that GST is in a way inhibited by OC. The MN test has been used as a measurement of genotoxicity in different animal groups including fish (Carrasco et al., 1990; Pacheco and Santos, 1998). Several types of pollutants, including PCB, DDT, PAH and metals, have shown to increase the MN fre quency in different fish species (reviewed in Al-Sabti and Metcalfe (1995)). One interesting result in this study is the significant lower levels of genotoxicity registered in wild seabream (0.3 ± 0.1) in com parison to the specimens from the aquaculture (from 1.0 to 2.5) showing that wild animals are not so exposed to pollutants that have the potential to induce genotoxic effects. Different fish spe cies can show different MN frequencies, nonetheless higher MN frequencies have been reported in more polluted sites (Minissi et al., 1996; Ergene et al., 2007), and after exposure to different types of pollutants (Bolognesi et al., 2006; Neuparth et al., 2006). The higher levels obtained in class III could be a direct result of a higher EROD activity. From the higher biotransformation enzyme activity could arise more metabol ites with genotoxic potential. Another hypothes is to be considered is the possible presence of more genotoxic chemicals in the land tanks the animals were transferred to, however the animals from the next weight class showed less MN. 5. Conclusions White seabream accumulates organochlorine compounds during the process of growth in the aquaculture production. Cultivated ani mals showed, in muscle and liver, higher values of lipids, tPCB and tDDT than wild animals. Biotransformation enzymes showed nega tive correlations with organochlorine levels in liver. M icronucleous numbers showed that wild white seabream are not so exposed to genotoxic compounds as cultured ones. In addition, biomarkers can be used also in terms of evaluating cultured fish quality versus wild. Contamin ant levels were low and comparable to other cultured spe cies giving further indications that white seabream is a species with a high potential for aquaculture production. Acknowledgements The authors would like to acknowledge the Portuguese Foun dation for Science and Technology (FCT) for the financial support
1673
(POCI/MAR/59094/2004). M. Ferreira, P. Antunes and J. Costa also thank the grants from the same institution. References Al-Sabti, K., Metcalfe, C.D., 1995. Fish micronuclei for assessing genotoxicity in water. Mutat. Res. 343, 121–135. Antunes, P., Gil, O., 2004. PCB and DDT contamination in cultivated and wild sea bass from Ria de Aveiro, Portugal. Chemosphere 54, 1503–1507. Antunes, P., Gil, O., Reis-Henriques, M.A., 2007. Evidence of higher biomagnification factors for lower chlorinated PCBs in cultivated seabass. Sci. Total Environ. 377, 36–44. Bolognesi, C., Perrone, E., Roggieri, P., Pampanin, D.M., Sciutto, A., 2006. Assessment of micronuclei induction in peripheral erythrocytes of fish exposed to xenobiot ics under controlled conditions. Aquat. Toxicol. 78S, S93–S98. Buckman, A.H., Wong, C.S., Chow, E.A., Brown, S.B., Solomon, K.R., Fisk, A.T., 2006. Biotransformation of polychlorinated biphenyls (PCBs) and bioformation of hydroxylated PCBs in fish. Aquat. Toxicol. 78, 176–185. Cachot, J., Geffard, O., Augagneur, S., Lacroix, S., Le Menach, K., Peluhet, L., Couteau, J., Denier, X., Devier, M.H., Pottier, D., Budzinski, H., 2006. Evidence of genotox icity related to high PAH content of sediments in the upper part of the Seine estuary (Normandy, France). Aquat. Toxicol. 79, 257–267. Carrasco, K.R., Tilbury, K.L., Myers, M.S., 1990. Assessment of the piscine micronu clei test as an in situ. Can. J. Fish. Aquat. Sci. 47, 2123–2136. Commandeur, J.N., Stijntjes, G.J., Vermeulen, N.P., 1995. Enzymes and transport sys tems involved in the formation and disposition of glutathione S-conjugates. Role in bioactivation and detoxication mechanisms of xenobiotics. Pharmacol. Rev. 47, 271–330. Couillard, C.M., Lebeuf, M., Ikonomou, M.G., Poirier, G.G., Cretney, W.J., 2005. Low hepatic ethoxyresorufin-O-deethylase activity correlates with high organochlo rine concentrations in Atlantic tomcod from the Canadian East Coast. Environ. Toxicol. Chem. 24, 2459–2469. Easton, M.D.L., Luszniak, D., Von der Geest, E., 2002. Preliminary examination of contaminant loadings in farmed salmon, wild salmon and commercial salmon feed. Chemosphere 46, 1053–1074. Ergene, S., Çavao, T., Çelik, A., Koleli, N., Kaya, F., Karahan, A., 2007. Monit oring of nuclear abnormalities in peripheral erythrocytes of three fish species from the Goksu Delta (Turkey): genotoxic damage in relation to water pollution. Ecotoxi cology 16, 385–391. Fent, K., Bucheli, T.D., 1994. Inhibition of hepatic microsomal monooxygenase sys tem by organotins in vitro in freshwater fish. Aquat. Toxicol. 28, 107–126. Fernandes, D., Porte, C., Bebianno, M.J., 2007. Chemical residues and biochemical responses in wild and cultured European sea bass (Dicentrarchus labra L.). Envi ron. Res. 103, 247–256. Ferreira, M., Antunes, P., Gil, O., Vale, C., Reis-Henriques, M.A., 2004. Organochlorine contaminants in flounder (Platichthys flesus) and mullet (Mugil cephalus) from Douro estuary, and their use as sentinel species for environmental monitoring. Aquat. Toxicol. 69, 347–357. Ferreira, M., Moradas-Ferreira, P., Reis-Henriques, M.A., 2006. The effect of longterm depuration on phase I and phase II biotransformation in mullets (Mugil cephalus) chronically exposed to pollutants in River Douro Estuary, Portugal. Mar. Environ. Res. 61, 326–338. Gallagher, E.P., Gross, T.S., Sheehy, K.M., 2001. Decreased glutathione S-transferase expression and activity and altered sex steroids in Lake Apopka brown bull heads (Ameiurus nebulosus). Aquat. Toxicol. 55, 223–237. Goksoyr, A., Forlin, L., 1992. The cytochrome P-450 system in fish, aquatic toxicology and environmental monitoring. Aquat. Toxicol. 22, 287–311. Habig, W.H., Pabst, M.J., Jakoby, W.B., 1974. Glutathione S-transferases – first enzy matic step in mercapturic acid formation. J. Biol. Chem. 249, 7130–7139. Jacobs, M.N., Covaci, A., Schepens, P., 2002. Investigation of selected persistent organic pollutants in farmed Atlantic salmon (Salmo salar), salmon aquaculture feed, and fish oil components of the feed. Environ. Sci. Technol. 36, 2797–2805. Kim, I.-Y., Hyun, C.-K., 2006. Comparat ive evaluation of the alkaline comet assay with the micronucleus test for genotoxicity monitoring using aquatic organ isms. Ecotox. Environ. Safe. 64, 288–297. Kunisue, T., Sakiyama, T., Yamada, T.K., Takahashi, S., Tanabe, S., 2007. Occurrence of hydroxylated polychlorinated biphenyls in the brain of cetaceans stranded along the Japanese coast. Mar. Pollut. Bull. 54, 963–973. Maule, A.G., Gannam, A.L., Davis, J.W., 2007. Chemical contaminants in fish feeds used in federal salmonid hatcheries in the USA. Chemosphere 67, 1308–1315. Minissi, S., Ciccotti, E., Rizzoni, M., 1996. Micronucleus test in erythrocytes of Barbus plebejus (Teleostei, Pisces) from two natural environments: a bioassay for the in situ detection of mutagens in freshwater. Mutat. Res. 367, 245–251. Neuparth, T., Bickham, J.W., Theod orakis, C.W., Costa, F.O., Costa, M.H., 2006. Endo sulfan-induced genotoxicity detected in the gilthead seabream, Sparus aurata L., by means of flow cytometry and micronuclei assays. Environ. Contam. Toxic ol. 76, 242–248. Pacheco, M., Santos, M.A., 1998. Induction of liver EROD and erythrocytic nuclear abnormalities by cyclophosphamide and PAHs in Anguilla anguilla L.. Ecotox. Environ. Safe. 40, 71–76. Perez, M.J., Rodriguez, C., Cejas, J.R., Martin, M.V., Jerez, S., Lorenzo, A., 2007. Lipid and fatty acid content in wild white seabream (Diplodus sargus) broodstock at different stages of the reproductive cycle. Comp. Biochem. Phys. B 146, 187– 196.
1674
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
Quental, T., Ferreira, A.M., Vale, C., 2003. The distribution of PCBs and DDTs in seston and plankton along the Portuguese coast. Acta Oecol. 24, S333– S339. Sa, R., Pousão-Ferreira, P., Oliva-Teles, A., 2006. Effect of dietary protein and lipid levels on growth and feed utilization of white sea bream (Diplodus sargus) juve niles. Aquacult. Nutr. 12, 310–321. Sa, R., Pousão-Ferreira, P., Oliva-Teles, A., 2007. Growth performance and metabolic utilization of diets with different protein: carbohydrate ratios by white sea bream (Diplodus sargus, L.) juveniles. Aquacult. Res. 38, 100–105. Saavedra, M., Conceição, L.E.C., Pousão-Ferreira, P., Dinis, M.T., 2006. Amino acid profiles of Diplodus sargus (L., 1758) larvae: implications for feed formulation. Aquaculture 261, 587–593. Singh, P.B., Singh, V., 2008a. Pesticide bioaccumulation and plasma sex steroids in fishes during the breeding phase from north India. Environ. Toxicol. Pharmacol. 25, 342–350.
Singh, P.B., Singh, V., 2008b. Bioaccumulation of hexachlorocyclohexane, dichlorodi phenyltrichloroethane, and estradiol-17b in catfish and carp during the premon soon season in India. Fish Physiol. Biochem. 34, 25–36. Toft, G., Edwards, T.M., Baatrup, E., Guillette Jr., L.J., 2003. Disturbed sexual character istics in male mosquitofish (Gambusia holbrooki) from a lake contaminated with endocrine disruptors. Environ. Health Persp. 111, 695–701. UNEP – United Nations Environment Programme, Secretariat of the Stockholm Con vention on Persistent Organic Pollutants, 2007. Guidance on the Global Monitor ing Plan for Persistent Organic Pollutants. Van der Oost, R., Beyer, J., Vermeulen, N.P.E., 2003. Fish bioaccumulation and bio markers in environmental risk assessment: a review. Environ. Toxic ol. Pharma col. 13, 57–149. Whyte, J.J., Jung, R.E., Schmitt, C.J., Tillitt, D.E., 2000. Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit. Rev. Toxic ol. 30, 347–570.
Contents lists available at ScienceDirect
Chemosphere j o u r n a l h o m e p a g e : w w w . e l s e v i e r. c o m / l o c a t e / ch e m o s p h e r e
Organochlorine bioaccumulation and biomarkers levels in culture and wild white seabream (Diplodus sargus) Marta Ferreira a,*, Paulo Antunes b,d, Joana Costa a, Joana Amado b, Odete Gil b, Pedro Pousão-Ferreira c, Carlos Vale b, Maria Armanda Reis-Henriques a,d a CIMAR/CIIMAR – Centro Interdisciplinar de Investigação Marinha e Ambiental, Universidade do Porto, Laboratório de Toxicologia Ambiental, Rua dos Bragas, 289, 4050-123 Porto, Portugal b INRB/IPIMAR – Instituto Nacional dos Recursos Biológicos, IPIMAR, Av. Brasília, 1449-006 Lisboa, Portugal c INRB/IPIMAR SUL – Instituto Nacional dos Recursos Biológicos, Av. 5 de Outubro, 8700-305 Olhão, Portugal d ICBAS/UP – Instituto de Ciências Biomédicas Abel Salaz ar, Universidade do Porto, Largo Professor Abel Salazar, 2, 4099-003 Porto, Portugal
a r t i c l e
i n f o
Article history: Received 24 March 2008 Received in revised form 13 June 2008 Accepted 24 July 2008 Available online 11 September 2008 Keywords: PCB DDT Fish farming EROD GST Micronucleous test
a b s t r a c t Persistent organic pollutants (POPs), which can accumulate in the adipose fish tissues, can enter the human food chain through the consumption of fish, and cause risk to health. The use of chemical analysis, and biochemical and cellular responses is a way to detect the impact of pollutants in aquatic systems. The purpose of this study was to investigate the levels of organochlorine compounds (polychlorinated biphe nyls – PCB and p,p9-dichlorodiphenyltrichloroethane and its metabol ites – tDDT) in, wild and cultivated, white seabream (Diplodus sargus), and also its biological effects that were evalua ted by assessing the activ ity of biotransformation enzymes and genotoxic effects. To achieve that we have sampled five different size classes (I – 13 g, II – 64 g, III – 143 g, IV – 315 g and V – 441 g) of white seabream from a local aquacul ture, and also a group of wild fish (375 g) in order to compare accumulation and responses between cul tured and wild fish. White seabream, cultured and wild, presented low levels of organochlorine content, both in liver and in muscle. Wild white seabream, in comparison to cultured ones at the marketable size, showed lower organochlorine accumulation. Biotransformation enzymes showed negative correlations with organochlorine levels in liver. Micronucleous numbers revealed that wild white seabream are not so exposed to genotoxic compounds as cultured ones. © 2008 Elsevier Ltd. All rights reserved.
1. Introduction In recent years, there has been an increasing awareness of the need to assess the adverse effects of contaminants in aquatic organisms, and also the risk of fish consumption for human health, mainly regarding farmed fish. Persistent organic pollutants (POPs), such as organochlorines, are lipophilic compounds that bioaccu mulate and biomagnify through the food chain (Singh and Singh, 2008a,b). The most widespread organochlor ines in the environ ment and in animal tissues are polychlorinated biphenyls (PCB), dichlorodiphenyltrichloroethane (DDT), and specially the DDT deg radation product, dichlorodiphenyldichloroethylene (DDE) (Toft et al., 2003), and their levels found in some marine organisms gave rise to some concern. In the marine environment, these hydropho bic compounds accumulate in the adipose tissues of fish thus enter ing the food chain. More recently, some concern has arise on halogenated com pounds in aquaculture systems namely the potential hazards * Corresponding author. Tel.: +351 22 340 18 00; fax: +351 22 339 06 08. E-mail address: mferreira@ciimar.up.pt (M. Ferreira). 0045-6535/$ - see front matter © 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2008.07.070
associated with the ingredients used in aquaculture feeds (Easton et al., 2002; Jacobs et al., 2002; Antunes and Gil, 2004; Maule et al., 2007). Moreover, several studies have reported higher orga nochlorine levels in cultivated species in comparison with wild species, like salmon (Easton et al., 2002) and seabass (Antunes and Gil, 2004). Among the available techniques, the integrated use of chemical analysis and biochemical and cellular responses is a way to detect the impact of anthropogenic contaminants in aquatic systems, both wild and cultivated fish (Ferreira et al., 2004; Fernandes et al., 2007). Xenobiotic compounds may be biotransformed in liver by enzymes from phase I and phase II. Phase I is a non-synthetic alter ation (oxidation, reduction or hydrolysis) of the original foreign molecule, which can then be conjugated in phase II (Commandeur et al., 1995). Biotransformation of lipophilic compounds is a requirement for detoxification and excretion (Ferreira et al., 2006). However, certain biotransformation steps are responsible for the activation of foreign chemicals to reactive intermediates that ultimately result in toxicity, genotoxicity, or carcinogenicity (Van der Oost et al., 2003). Organochlorine compounds are potentially genotoxic, implying that they can damage DNA directly and/or
1670
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
enerate reactive species which can in turn damage DNA (Cachot g et al., 2006). The micronucleous (MN) test in fish has been shown to be a useful in vivo technique for genotoxicity testing and to have potential for in situ monitoring of water quality; the MN test detects micronuclei resulting from either chromosomal breakage during cell division or chromosome loss events in anaphase dam ages (Kim and Hyun, 2006). The marine aquaculture in the South-European countries is based on gilthead seabream (Sparus aurata) and seabass (Dicentra chus labrax) (Saavedra et al., 2006). However, the introduction of new species in aquaculture is an issue in development, and the white seabream (Diplod us sargus) is considered to be a promising new species with high market values and demand (Sa et al., 2006, 2007). The white seabream is an omnivorous species of the Sparidae family, found in the Mediterranean as well as in the eastern Atlantic Ocean, including the archipelagos of Madeira, Cape Verde and the Canary Islands (reviewed in Perez et al. (2007)). The purpose of this study was to investigate the levels of orga nochlorine compounds (tPCB and tDDT) in cultivated white sea bream (in different growth stages) and in the corresponding food pellets, and also the biological effects. The white seabream was chosen because of its high potential for aquaculture. This species is found in the Mediterranean, as well as in the Atlantic Ocean, and is mainly caught by self-employed fisherman and constitutes a valu able fishery resource considering its high price (Perez et al., 2007). The biological effects were evaluated by assessing the phase I and II biotransformation enzymes (EROD and GST, respectively) and genotoxic effects through the MN test. Identical study was also con ducted with the wild specim ens of white seabream. Moreover, to our knowledge the evaluation of the biotransformation enzymes activit ies, EROD and GST, widely used as biomarkers of exposure to pollutants, and the MN test to assess genotoxic effects, has never been reported in this particular species. 2. Materials and methods
weighted and the condition factor (CF) calculated (body weight (g) £ 100/length3 (cm)). Blood was collected from the caudal vein to the MN test. Liver and muscle were sampled for biochemical analysis and measurement of organochlorine compounds levels. The samples were transported to the laboratory in dry ice; samples for biochemical analysis were stored at ¡80 °C and for organochlo rine determinations at ¡20 °C. Water from the land tanks was filtered through pre-washed (hexane) and pre-combusted (350 °C, 17 h) Gelman A/E filters and particulate fraction collected. Filters were stored frozen and then dried at 40 °C for analysis. 2.2. PCB and DDT analysis 18 PCB congeners (IUPAC Nos. 18, 28, 52, 49, 44, 101, 151, 149, 118, 153, 105, 138, 187, 183, 128, 180, 170, 194), p,p9-DDT and metab olites (p,p9-DDD and p,p9-DDE) were quantified. Analyses were per formed according to Antunes and Gil (2004). Values are presented as the sum of the 18 congeners – tPCB, the sum of 7 indicator PCB namely PCB28, 52, 101, 118, 138, 153 and 180 – PCB7 (UNEP, 2007), and the sum of p,p9-DDT and its metabolites – tDDT. About 200 g of diet pellets was collected and homogenized. Subsamples of 2.00 g of diet pellets and muscle and 0.5000 g of liver were Soxhlet extracted with n-hexane for 6 h, and suspended particulate matter (SPM) samples were also Soxhlet extracted for 16 h. Lipid content was determined gravimetrically from aliquots of tissue extracts. The remaining extract was purified with a Florisil column and further with sulphuric acid before the analysis in an Agilent 6890N gas chromatograph, equipped with a micro elec tron capture detector and a DB-5 (J&W Scientific) capillary column (60 m £ 0.25 mm i.d. £ 0.25 lm film thickness). tPCB and tDDT were quantified using a six point calibration curve and CBs 65 and 204 as internal standards. Procedural blanks were analysed each 10–16 samples to monitor possible laboratory contamination. Detection limits calculated from three times the peak height in blank sam ples, ranged from 0.01 to 0.04 ng g ¡1 dw.
2.1. Sampling 2.3. Biochemical analysis White seabream samples were collected from a fish farm located in Ria Formosa, Olhão, in the South of Portugal, and wild specimens were captured in the coastal zone. The wild specimens sampled were correspondent to the higher weight classes, the marketable size, in order to compare the fish quality (wild and cul tured) for human consumption. Five different size classes (Table 1) according to their weight and a sub-sample of the food pellet correspondent to each size class were sampled. The two smaller classes were sampled from glass fibber tanks, with constant filtered water circulation, and the larger specimens from the land tanks with water circulation con trolled by the tidal cycle. The fish were killed by asphyxia in ice, and the tissues sampled at the fish farm laboratory. Wild animals were captured by anglers. The animals were immediately killed by asphyxia in ice and sampled in place. Animals were measured and
Table 1 Weight (g), length (cm) and condition factor (CF) for the different size classes of white seabream Class
N
Weight (g)
Length (cm)
CF
I II III IV V Wild
24 24 10 6 8 7
12.8 ± 0.3 63.7 ± 3.2 142.7 ± 5.8 315.5 ± 8.3 441.4 ± 11.9 375.3 ± 41.9
8.6 ± 0.1 14.8 ± 0.2 20.6 ± 0.4 25.2 ± 0.3 28.8 ± 0.3 27.6 ± 1.2
1.97 ± 0.03a 1.95 ± 0.03a,b 1.65 ± 0.05c 1.97 ± 0.10a,d 1.85 ± 0.03b,d,e 1.75 ± 0.05c,e
Values presented as mean ± SE. Different letters denote significant differences among the different groups, p < 0.05.
Livers were homogenized in ice-cold sodium phosphate buffer 50 mM, Na2EDTA 0.1 mM, pH 7.8 and centrifuged at 15000g for 20 min, at 4 °C. Glutathione S-transferase (GST) in the liver was determined according to the method of Habig et al. (1974) adapted to microplate as described in Ferreira et al. (2006) using glutathi one (GSH) 10 mM in phosphate buffer 0.1 M, pH 6.5, and 1-chloro2,4-dinitrobenzene (CDNB) 60 mM in ethanol prepared just before the assay. The reaction mixture consisted of phosphate buffer, GSH solution and CDNB solution in a proportion of 4.95 mL (phosphate buffer): 0.9 mL (GSH): 0.15 mL (CDNB). In the microplate, 0.2 mL of the reaction mixture was added to 0.1 mL of the sample, with final concentration 1 mM GSH and 1 mM CDNB in the assay. The GST activity was measured immediately every 20 s, at 340 nm, dur ing the first 5 min, and calculated in the period of linear change in absorbance. Liver GSH activity is expressed in nmol/min/mg protein. Liver ethoxyresorufin O-deethylase (EROD) activity was mea sured according to Ferreira et al. (2004). Briefly, liver was homog enized in ice-cold buffer (50 mM Tris–HCl, pH 7.4, 0.15 M KCl). Microsomes were obtained by centrifugation of the 9000g super natant at 36000g for 90 min in a SIGMA 3K30 centrifuge. The pellet was then resuspended in buffer (50 mM Tris–HCl, 1 mM NaEDTA, pH 7.4, 1 mM dithiothreitol, 20% v/v glycerol) and spun down at 36000g for 120 min (Fent and Bucheli, 1994). Microsomes were suspended in EDTA-free resuspension buffer and stored at ¡80 °C until use. Microsomal suspension (50 lL) was incubated with eth oxyresorufin 0.5 lM for 1 min, and the enzymatic reaction was
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
initiated by the addition of 45 lM NADPH. EROD activity was mea sured for 5 min at kex 530 nm and kem 585 nm, and determined by comparison to a resorufin standard curve. Hepatic EROD activity is expressed in pmol/min/mg protein. 2.4. Micronuclei test Two blood smear slides were prepared for each fish, fixated in methanol for 10 min and stained with Giemsa 5% in 3 mM phos phate buffer for 30 min. Two slides per fish were observed under a light microscope and micronuclei were recorded in a total of 1000 erythrocytes per slide. 2.5. Statistical analysis Differences between groups were tested using a One-Way ANOVA with a multiple comparison test (LSD) at a 5% significance level. Some data had to be log transformed in order to fit ANOVA assumptions. All tests were performed using the software Statis tica 6.0 (Statsoft, Inc., 2001). 3. Results The concentrations of organochlorine compounds in the food pellets and in suspended particulate matter (SPM) from the correspondent pounds are shown in Table 2. Values of the seven indicator PCB (PCB7) are also presented in Tables 2 and 3 to allow comparisons with other studies, the PCB7 represent about 65% of tPCB without significant differences to tPCB, therefore results will be discussed as tPCB. SPM showed low levels of tPCB and tDDT. The food pellets to feed animals from class III had the highest concentrations of these two types of organochlorine compounds. Table 2 Lipid content (%), sum of all analyzed PCB congeners – tPCB, sum of 7 indicator PCB congeners – PCB7, and sum of p,p9-DDT and its metabolites – tDDT, in the food pellets (ng g¡1 lipids) from each class of white seabream, and in the suspended par ticulate matter (SPM) (ng g¡1) Food pellet Classes I + II Class III Classes IV + V
Lipids (%) 19.4 16.9 8.9
tPCB ng g¡1 lipids 45 150 108
PCB7 ng g¡1 lipids 23 83 57
tDDT ng g¡1 lipids 26 45 21
SPM Class III Class IV
– –
ng g¡1 11 8
ng g¡1 5.4 3.7
ng g¡1 1.1 0.6
1671
The pellets of classes I + II presented the lowest levels of tPCB, how ever, the concentrations of tDDT were not different to the pellets supplied to fish from classes IV + V. Levels of lipid content, tPCB and tDDT in muscle and liver of cultivated and wild white seabream are displayed in Table 3. Ani mals from class I presented higher levels of lipid content in liver in comparison to the other classes although difference to class II and V were not significant. Lipid content in liver was not higher than in muscle, with the exception of individuals of class I. Regarding the organochlorine compounds, tPCB levels were always higher than tDDT. In white seabream from class II and wild animals tPCB levels were significantly higher in liver, in contrast with tDDT that were similar in both tissues. Muscle of individuals of class III showed higher levels of organochlorines than liver, and in the other size classes the two tissues showed no significant differences. Fish’s muscle from class III showed the higher content of tPCB and tDDT in comparison with the other analysed classes. Comparing wild animals with cultivated ones, of similar sizes (class IV and V), we can observe that wild seabream showed, in muscle and in liver, lower values of lipids, tPCB and tDDT than cul tivated animals (Table 3). In both tissues p,p9-DDT represented less than 27% of tDDT in farmed white seabream, and less than 35% in wild fish. Biomarkers evaluated in this study, hepatic EROD and GST activ ity, and MN in erythrocytes are presented in Table 4. White sea bream from class III showed significant higher levels of hepatic EROD activity, in comparison to the other classes analysed. This result is in agreement with the higher levels of organochlorine con taminants in the muscle. The fact that these fish had recently been transferred to the land tanks and been in the estuary waters could have influenced this biomarker. The lowest value for EROD activ ity was registered in animals from class V, which also showed the higher accumulation of tPCB and tDDT in liver. The wild specimens, that presented considerable lower levels of organochlorine accu mulation, showed EROD activity values similar to the ones from class II and IV. Overall, a significant negative correlation was found between tPCB and tDDT levels in liver with hepatic EROD activity. The accumulation of contaminants in liver and in muscle reflects different inputs; in muscle chronic accumulation is reflected while liver represents recent inputs. With this in mind we have correlated phase I enzyme with tPCB accumulation in muscle (Fig. 1). Inter estingly, a positive correlation was found for this enzyme activity and the tPCB accumulation in muscle; animals with lower weight showed higher levels of hepatic EROD activity and lower levels of tPCB in muscle; on the contrary, animals with higher weight reflected more tPCB accumulation and lower EROD activity.
Table 3 Lipid content (%), sum of all analyzed PCB congeners – tPCB, sum of 7 indicator PCB congeners – PCB7, and sum of p,p9-DDT and its metabolites – tDDT, in liver and muscle of white seabream in ng g¡1 lipids Class
n
I
6
II
6
III
10
IV
6
V
8
Wild
7
Lipids (%)
tPCB (ng g¡1 lipids)
PCB7 (ng g¡1 lipids)
tDDT (ng g¡1 lipids)
Liver
Muscle
Liver
Muscle
Liver
Muscle
Liver
Muscle
37.7 ± 6.7a (8.2–57.7) 27.1 ± 3.0a,b (18.3–38.8) 16.7 ± 1.9b,c (7.2–25.3) 14.3 ± 1.1c,d (9.7–17.4) 29.2 ± 3.3a (15.3–39.5) 9.7 ± 0.6d (8.1–12.2)
12.3 ± 0.9a,c (7.9–14.0) 20.5 ± 0.7b,d (18.4–20.9) 11.4 ± 1.6a,c (5.2–18.6) 19.3 ± 2.6a,b (7.6–25.8) 27.1 ± 2.3d (16.4–35.3) 8.5 ± 1.6c (2.9–13.9)
117.3 ± 4 4.5a (62.8–339.2) 212.6 ± 7.9a,b (185.5–242.2) 325.7 ± 56.6b (118.2–729.0) 297.1 ± 28.5b (191.4–380.8) 529.0 ± 42.5c (408.0–675.1) 156.8 ± 13.8a (116.7–215.6)
101.5 ± 4.8a (87.0–117.8) 167.2 ± 4.7b (153.2–183.1) 736.4 ± 66.4c (504.5–1016.3) 430.3 ± 70.3d (261.2–760.5) 471.4 ± 30.0d (358.5–649.0) 111.3 ± 12.6a (73.3–163.3)
78.3 ± 30.1a (40.3–228.0) 124.2 ± 3.6a,b (116.0–137.3) 222.9 ± 47.9c (81.5–482.5) 202.4 ± 20.1b,c (265.3–435.5) 338.3 ± 25.4d (265.3–435.5) 89.9 ± 7.9a (68.8–125.8)
57.9 ± 2.6a (51.3–66.9) 105.9 ± 3.4b (94.5–117.2) 500.0 ± 47.8c (315.9–698.7) 301.4 ± 49.8d (189.7–536.7) 323.8 ± 20.9d (243.0–443.7) 70.6 ± 9.1a (45.5–111.6)
59.3 ± 21.1a,b (30.4–164.1) 45.5 ± 2.8a,c (34.3–54.3) 42.2 ± 8.3a,c (7.7–86.0) 38.7 ± 6.3a,c (24.1–65.7) 77.8 ± 10.9b (44.4–134.7) 24.4 ± 2.0c (17.3–34.6)
34.0 ± 1.6a.b (28.1–39.1) 52.9 ± 2.0a,c (44.8–58.8) 103.2 ± 15.4d (52.8–66.4) 59.8 ± 9.9a,c (28.3–96.9) 71.7 ± 9.8c (45.6–124.5) 19.8 ± 9.8b (14.5–29.1)
Values presented as mean ± SE (minimum–maximum). Different letters denote significant differences among the different groups, p < 0.05.
1672
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
Table 4 Hepatic EROD and GST activities, and erythrocyte micronucleus (MN) in white sea bream Class
n
EROD (pmol/min/ mg protein)
GST (nmol/min/ mg protein)
MN (per 1000 erythrocytes)
I II III IV V Wild
6 6 10 6 8 7
80.0 ± 10.6a,c 48.3 ± 14.5a,b,d 125.1 ± 23.8c 44.0 ± 14.9b,d 21.6 ± 3.0d 55.1 ± 7.4a,b
104.2 ± 4.3a 108.4 ± 10.1a,b 118.8 ± 5.4a,b 127.2 ± 7.3b,c 76.8 ± 3.2d 148.8 ± 8.7c
2.1 ± 0.4a 2.0 ± 0.5a 2.5 ± 0.4a 1.0 ± 0.5b,c 1.9 ± 0.5a,b 0.3 ± 0.1c
Different letters denote significant differences among the different groups, p < 0.05.
Fig. 1. Correlation between hepatic EROD activity (pmol/min/mg protein) and tPCB (ng g¡1 lipids) in muscle in the lower classes (r) and in the higher weight classes (I).
Regarding hepatic GST activity the values are presented in Table 4, an increase in activity was observed from class I to IV. On the con trary, and in agreement with EROD activity, class V has also showed significant lower levels for this enzyme. The wild specimens pre sented significant higher levels for this biomarker. As for EROD activity a negative significant correlation was found between GST and tPCB in liver (r = 0.67; p < 0.05), and also with tDDT (r = 0.63; p < 0.05). The genotoxicity biomarker, evaluated as MN numbers, did not show significant differences between the evaluated classes, with the exception of class IV that showed lower values than the smaller classes of white seabream; nevertheless, class III has also presented a slightly higher value, as for EROD activity. Interestingly, and con trary to the other biomarkers assessed, wild specimens of white seabream presented significant lower numbers of MN, showing that wild animals are not so exposed to contaminants with geno toxic properties as animals from aquaculture. 4. Discussion The aim of the present study was to evaluate the levels of orga nochlorine contaminants (tPCB and tDDT) in a cultured species, the white seabream (D. sargus); and to detect and assess the possible adverse effects of these contaminants at a biochemical and cellular level, in different life stages of the species. In addition, to compare cultured specimens with wild specimens captured offshore. The observed differences between wild and cultivated fish may be explained by the different environments that result in different condition factors (Table 1). Cultivated fish have higher condition factor, as a result of the abundant food supply and closed environ ment. If we compare wild seabream only with the classes IV and V, both with a marketable size, the wild fish presented a lower
CF, the same has been observed for seabass wild and cultured, with wild seabass presenting significant lower CF (Fernandes et al., 2007). The fish farm is also located in a coastal lagoon, which contains higher levels of organic matter and contaminants than coastal waters (Quental et al., 2003). The higher levels found in farmed fish do not represent any risk to human consumption and are in the same order of other farmed and estuarine fish (Antunes and Gil, 2004). Excluding muscle of class III fish, tPCB levels presented an increase with fish size/age in both analysed tissues, and tDDT levels showed small variations. The fact that fish from class III presented lower lipid contents and higher organo chlorine levels than the other size classes is an unexplained result. One possible explanation is the higher content of organochlorine compounds in the food pellets supplied to the fish of this class (Table 2). However, the fact that this increase was not observed in liver make us consider other possibilities: (i) these fish were recently moved from indoor fibber glass tanks to outdoor land tanks, which could produce a stress factor that lead to an internal redistribution of contaminants in this adaptation period, this is in agreement with the lower condition factor found in this size class; (ii) there may be some differences in contaminant exposure due to sediment and suspended particulate matter; or (iii) there may occur variations in metabolic capacities of fish at this grow ing stage. Similar findings were obtained in another species, sea bass (D. labrax), produced in a similar fish farm (Antunes et al., 2007), with animals that still were in fibber glass tanks. This rein forces the idea of these variations being a normal biologic effect of fish growth. Indeed, these changes in metabolic capacities could explain higher hepatic EROD activity in white seabream class III, as described in Section 3. Biological effects in white seabream were evaluated by means of EROD and GST activity, and MN frequency. The referred biomark ers are not specific for PCB and DDT; however the integrated use of chemical and biological analysis is a way to detect the effects of exposure to contaminants (van der Oost et al., 2003; Ferreira et al., 2006). Phase I biotransformation enzyme, evaluated in this study by means of hepatic EROD activity has been described to be involved in the metabolic elimination of several contaminants, like PAH, PCB and others (Goksoyr and Forlin, 1992; Whyte et al., 2000; Ferreira et al., 2004, 2006). Liver is the organ where biotransforma tion of contaminants is processed; however we have found that ani mals presenting higher levels of OC accumulation, like the ones in class IV and V showed lower levels of EROD activity. In agreement with this are the wild animals that showed significant lower lev els of OC in liver and higher values for this enzyme. Same results were obtained in Atlantic tomcod from the Canadian coast, where the authors have found low hepatic EROD activity in correlation with higher levels of organochlorine contaminants (Couillard et al., 2005). Some other studies performed with largemouth bass, in a contaminated site with PCB has shown a moderate induction of EROD activity, and that after a short period of time enzyme levels have declined, suggesting a catalytic inhibition of this enzyme by certain PCB congeners (reviewed in Whyte et al. (2000)). Another possible explanation for the higher activity in smaller fish is that, usually smaller fish have a higher metabolic activity per gram (Whyte et al., 2000). Finally, we must not exclude the fact that hydroxylated PCB have been found in the environment and also in fish tissues (Buckman et al., 2006; Kunisue et al., 2007), indicat ing biotransformation from CYP enzyme-mediated, however the suggestion is that they could be biotransformed not by CYP1A but by CYP2B (Buckman et al., 2006), not measurable by hepatic EROD activity. One interesting results is the significant elevated EROD activity in class III. This could be a direct result of the transfer of the fish from the fiber glass tanks, with filtered water, to the land tanks with estuarine water that could have more EROD inducers present, like PAHs well documented EROD inducers (Van der Oost
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
et al., 2003; Ferreira et al., 2006). On the contrary, EROD activity showed a significant positive correlation with PCB accumulation in muscle (Fig. 1). Considering the muscle concentrations as a better indicator of total contamination in fish, adult white seabream pre sented a lower EROD activity response than juvenile. Taking into account only adults of commercial size, we obtained a positive cor relation between the hepatic EROD activity and PCB levels in the edible tissue, showing that EROD activity can be considered a good biomarker for total PCB contamination in fish. Glutathione S-transferase (GST) was assessed as phase II in the metabolic process of the biotransformation of contaminants. With exception of the fish from class V that presented significant lower levels of activity, the other weight classes presented sim ilar values between them. Overall, it was observed a significant negative correlation between this enzymes and the accumulation of OC. The decrease in this enzyme activity, in the presence of organochlorine contamin ants, has already been reported. A study performed with mullets has reported that before depuration, and with higher content in tPCB and tDDT, GST presented lower lev els of activity (Ferreira et al., 2006). Gallagher et al. (2001) have also stated that in brown bullhead, PCB have the ability to reduce GST expression. In fact the significant higher levels of GST activ ity presented by the wild white seabream than farmed seabass with simil ar weight (class IV and V), that showed considerable lower levels of OC accumulation can suggest that GST is in a way inhibited by OC. The MN test has been used as a measurement of genotoxicity in different animal groups including fish (Carrasco et al., 1990; Pacheco and Santos, 1998). Several types of pollutants, including PCB, DDT, PAH and metals, have shown to increase the MN fre quency in different fish species (reviewed in Al-Sabti and Metcalfe (1995)). One interesting result in this study is the significant lower levels of genotoxicity registered in wild seabream (0.3 ± 0.1) in com parison to the specimens from the aquaculture (from 1.0 to 2.5) showing that wild animals are not so exposed to pollutants that have the potential to induce genotoxic effects. Different fish spe cies can show different MN frequencies, nonetheless higher MN frequencies have been reported in more polluted sites (Minissi et al., 1996; Ergene et al., 2007), and after exposure to different types of pollutants (Bolognesi et al., 2006; Neuparth et al., 2006). The higher levels obtained in class III could be a direct result of a higher EROD activity. From the higher biotransformation enzyme activity could arise more metabol ites with genotoxic potential. Another hypothes is to be considered is the possible presence of more genotoxic chemicals in the land tanks the animals were transferred to, however the animals from the next weight class showed less MN. 5. Conclusions White seabream accumulates organochlorine compounds during the process of growth in the aquaculture production. Cultivated ani mals showed, in muscle and liver, higher values of lipids, tPCB and tDDT than wild animals. Biotransformation enzymes showed nega tive correlations with organochlorine levels in liver. M icronucleous numbers showed that wild white seabream are not so exposed to genotoxic compounds as cultured ones. In addition, biomarkers can be used also in terms of evaluating cultured fish quality versus wild. Contamin ant levels were low and comparable to other cultured spe cies giving further indications that white seabream is a species with a high potential for aquaculture production. Acknowledgements The authors would like to acknowledge the Portuguese Foun dation for Science and Technology (FCT) for the financial support
1673
(POCI/MAR/59094/2004). M. Ferreira, P. Antunes and J. Costa also thank the grants from the same institution. References Al-Sabti, K., Metcalfe, C.D., 1995. Fish micronuclei for assessing genotoxicity in water. Mutat. Res. 343, 121–135. Antunes, P., Gil, O., 2004. PCB and DDT contamination in cultivated and wild sea bass from Ria de Aveiro, Portugal. Chemosphere 54, 1503–1507. Antunes, P., Gil, O., Reis-Henriques, M.A., 2007. Evidence of higher biomagnification factors for lower chlorinated PCBs in cultivated seabass. Sci. Total Environ. 377, 36–44. Bolognesi, C., Perrone, E., Roggieri, P., Pampanin, D.M., Sciutto, A., 2006. Assessment of micronuclei induction in peripheral erythrocytes of fish exposed to xenobiot ics under controlled conditions. Aquat. Toxicol. 78S, S93–S98. Buckman, A.H., Wong, C.S., Chow, E.A., Brown, S.B., Solomon, K.R., Fisk, A.T., 2006. Biotransformation of polychlorinated biphenyls (PCBs) and bioformation of hydroxylated PCBs in fish. Aquat. Toxicol. 78, 176–185. Cachot, J., Geffard, O., Augagneur, S., Lacroix, S., Le Menach, K., Peluhet, L., Couteau, J., Denier, X., Devier, M.H., Pottier, D., Budzinski, H., 2006. Evidence of genotox icity related to high PAH content of sediments in the upper part of the Seine estuary (Normandy, France). Aquat. Toxicol. 79, 257–267. Carrasco, K.R., Tilbury, K.L., Myers, M.S., 1990. Assessment of the piscine micronu clei test as an in situ. Can. J. Fish. Aquat. Sci. 47, 2123–2136. Commandeur, J.N., Stijntjes, G.J., Vermeulen, N.P., 1995. Enzymes and transport sys tems involved in the formation and disposition of glutathione S-conjugates. Role in bioactivation and detoxication mechanisms of xenobiotics. Pharmacol. Rev. 47, 271–330. Couillard, C.M., Lebeuf, M., Ikonomou, M.G., Poirier, G.G., Cretney, W.J., 2005. Low hepatic ethoxyresorufin-O-deethylase activity correlates with high organochlo rine concentrations in Atlantic tomcod from the Canadian East Coast. Environ. Toxicol. Chem. 24, 2459–2469. Easton, M.D.L., Luszniak, D., Von der Geest, E., 2002. Preliminary examination of contaminant loadings in farmed salmon, wild salmon and commercial salmon feed. Chemosphere 46, 1053–1074. Ergene, S., Çavao, T., Çelik, A., Koleli, N., Kaya, F., Karahan, A., 2007. Monit oring of nuclear abnormalities in peripheral erythrocytes of three fish species from the Goksu Delta (Turkey): genotoxic damage in relation to water pollution. Ecotoxi cology 16, 385–391. Fent, K., Bucheli, T.D., 1994. Inhibition of hepatic microsomal monooxygenase sys tem by organotins in vitro in freshwater fish. Aquat. Toxicol. 28, 107–126. Fernandes, D., Porte, C., Bebianno, M.J., 2007. Chemical residues and biochemical responses in wild and cultured European sea bass (Dicentrarchus labra L.). Envi ron. Res. 103, 247–256. Ferreira, M., Antunes, P., Gil, O., Vale, C., Reis-Henriques, M.A., 2004. Organochlorine contaminants in flounder (Platichthys flesus) and mullet (Mugil cephalus) from Douro estuary, and their use as sentinel species for environmental monitoring. Aquat. Toxicol. 69, 347–357. Ferreira, M., Moradas-Ferreira, P., Reis-Henriques, M.A., 2006. The effect of longterm depuration on phase I and phase II biotransformation in mullets (Mugil cephalus) chronically exposed to pollutants in River Douro Estuary, Portugal. Mar. Environ. Res. 61, 326–338. Gallagher, E.P., Gross, T.S., Sheehy, K.M., 2001. Decreased glutathione S-transferase expression and activity and altered sex steroids in Lake Apopka brown bull heads (Ameiurus nebulosus). Aquat. Toxicol. 55, 223–237. Goksoyr, A., Forlin, L., 1992. The cytochrome P-450 system in fish, aquatic toxicology and environmental monitoring. Aquat. Toxicol. 22, 287–311. Habig, W.H., Pabst, M.J., Jakoby, W.B., 1974. Glutathione S-transferases – first enzy matic step in mercapturic acid formation. J. Biol. Chem. 249, 7130–7139. Jacobs, M.N., Covaci, A., Schepens, P., 2002. Investigation of selected persistent organic pollutants in farmed Atlantic salmon (Salmo salar), salmon aquaculture feed, and fish oil components of the feed. Environ. Sci. Technol. 36, 2797–2805. Kim, I.-Y., Hyun, C.-K., 2006. Comparat ive evaluation of the alkaline comet assay with the micronucleus test for genotoxicity monitoring using aquatic organ isms. Ecotox. Environ. Safe. 64, 288–297. Kunisue, T., Sakiyama, T., Yamada, T.K., Takahashi, S., Tanabe, S., 2007. Occurrence of hydroxylated polychlorinated biphenyls in the brain of cetaceans stranded along the Japanese coast. Mar. Pollut. Bull. 54, 963–973. Maule, A.G., Gannam, A.L., Davis, J.W., 2007. Chemical contaminants in fish feeds used in federal salmonid hatcheries in the USA. Chemosphere 67, 1308–1315. Minissi, S., Ciccotti, E., Rizzoni, M., 1996. Micronucleus test in erythrocytes of Barbus plebejus (Teleostei, Pisces) from two natural environments: a bioassay for the in situ detection of mutagens in freshwater. Mutat. Res. 367, 245–251. Neuparth, T., Bickham, J.W., Theod orakis, C.W., Costa, F.O., Costa, M.H., 2006. Endo sulfan-induced genotoxicity detected in the gilthead seabream, Sparus aurata L., by means of flow cytometry and micronuclei assays. Environ. Contam. Toxic ol. 76, 242–248. Pacheco, M., Santos, M.A., 1998. Induction of liver EROD and erythrocytic nuclear abnormalities by cyclophosphamide and PAHs in Anguilla anguilla L.. Ecotox. Environ. Safe. 40, 71–76. Perez, M.J., Rodriguez, C., Cejas, J.R., Martin, M.V., Jerez, S., Lorenzo, A., 2007. Lipid and fatty acid content in wild white seabream (Diplodus sargus) broodstock at different stages of the reproductive cycle. Comp. Biochem. Phys. B 146, 187– 196.
1674
M. Ferreira et al. / Chemosphere 73 (2008) 1669–1674
Quental, T., Ferreira, A.M., Vale, C., 2003. The distribution of PCBs and DDTs in seston and plankton along the Portuguese coast. Acta Oecol. 24, S333– S339. Sa, R., Pousão-Ferreira, P., Oliva-Teles, A., 2006. Effect of dietary protein and lipid levels on growth and feed utilization of white sea bream (Diplodus sargus) juve niles. Aquacult. Nutr. 12, 310–321. Sa, R., Pousão-Ferreira, P., Oliva-Teles, A., 2007. Growth performance and metabolic utilization of diets with different protein: carbohydrate ratios by white sea bream (Diplodus sargus, L.) juveniles. Aquacult. Res. 38, 100–105. Saavedra, M., Conceição, L.E.C., Pousão-Ferreira, P., Dinis, M.T., 2006. Amino acid profiles of Diplodus sargus (L., 1758) larvae: implications for feed formulation. Aquaculture 261, 587–593. Singh, P.B., Singh, V., 2008a. Pesticide bioaccumulation and plasma sex steroids in fishes during the breeding phase from north India. Environ. Toxicol. Pharmacol. 25, 342–350.
Singh, P.B., Singh, V., 2008b. Bioaccumulation of hexachlorocyclohexane, dichlorodi phenyltrichloroethane, and estradiol-17b in catfish and carp during the premon soon season in India. Fish Physiol. Biochem. 34, 25–36. Toft, G., Edwards, T.M., Baatrup, E., Guillette Jr., L.J., 2003. Disturbed sexual character istics in male mosquitofish (Gambusia holbrooki) from a lake contaminated with endocrine disruptors. Environ. Health Persp. 111, 695–701. UNEP – United Nations Environment Programme, Secretariat of the Stockholm Con vention on Persistent Organic Pollutants, 2007. Guidance on the Global Monitor ing Plan for Persistent Organic Pollutants. Van der Oost, R., Beyer, J., Vermeulen, N.P.E., 2003. Fish bioaccumulation and bio markers in environmental risk assessment: a review. Environ. Toxic ol. Pharma col. 13, 57–149. Whyte, J.J., Jung, R.E., Schmitt, C.J., Tillitt, D.E., 2000. Ethoxyresorufin-O-deethylase (EROD) activity in fish as a biomarker of chemical exposure. Crit. Rev. Toxic ol. 30, 347–570.