Assessment of the pollution impact on biomarkers of effect of a freshwater fish

Chemosphere 68 (2007) 1582–1590 www.elsevier.com/locate/chemosphere

Assessment of the pollution impact on biomarkers of effect of a freshwater fish Fernando R. de la Torre a

a,b,*

, Alfredo Salibia´n

a,c

, Lucrecia Ferrari

a,c

Applied Ecophysiology Program, Basic Sciences Department, National University of Lujan, Casilla de Correo 221, B6700ZBA Luja´n, Argentina b National Scientific and Technical Research Council (CONICET), Argentina c Scientific Research Commision (CIC), La Plata, Buenos Aires, Argentina Received 23 June 2006; received in revised form 23 December 2006; accepted 18 February 2007 Available online 16 April 2007

Abstract Changes in biomarkers of fish captured from stressed environments may represent a reliable tool in revealing sublethal effects of the pollutants found in aquatic ecosystems. The response patterns of selected biochemical and morphological variables biomarkers of effect were assessed in adult females of an indigenous teleost Cnesterodon decemmaculatus, caught at a polluted site (San Francisco) of the Reconquista river (Buenos Aires, Argentina). Combined field-caging experiments with controlled laboratory exposure to clean media were performed. The biochemical parameters measured were specific activities of gill (Na+ + K+)-ATPase, liver AlaAT and AspAT, and brain AChE; LSI and CF were also calculated. The changes in gill (Na+ + K+)-ATPase and liver AlaAT activities of fish captured in the field in most cases were reversible after transfer to clean media. The results were interpreted in association with the physicochemical profile of the water samples taken simultaneously with the capture of the fish. Results suggested the suitability of the test species used as tools in environmental monitoring programs of risk assessment.  2007 Elsevier Ltd. All rights reserved. Keywords: Freshwater toxicity assessment; Fish biomarkers; Brain acetylcholinesterase; Gill (Na+ + K+)-ATPase; Liver aminotransferases; Reversibility of altered fish biomarkers; Cnesterodon decemmaculatus

1. Introduction In addition to classical studies of freshwater pollution based on media physicochemical analyses, efforts are being made to identify and design new approaches and tools for diagnosing water toxicity. Analysis of chemical substances in tissues and body fluids, toxic metabolites, enzymes activities and other biochemical variables have frequently been used in documenting the toxin interaction with biological systems. Changes detected in substance concentration measurements, known as toxicity biomarkers, are recognized

* Corresponding author. Address: Applied Ecophysiology Program, Basic Sciences Department, National University of Lujan, Casilla de Correo 221, B6700ZBA Luja´n, Argentina. E-mail address: [email protected] (F.R. de la Torre).

0045-6535/$ - see front matter  2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.02.033

that links the effects of contaminant exposure and risk assessment processes. The term biomarker used in a broad sense includes measuring parameters providing the means to reflect the interaction between a particular biological structure and stressful factors found in the environment under consideration, which may be chemical, physical or biological (Lagadic et al., 1997; van der Oost et al., 2003). According to WHO (1993) biomarkers can be grouped into three general categories: biomarkers of exposure, of effect and of susceptibility. Since they give information on the biological effects of pollutants rather than a quantification of their environmental levels, we studied the physicochemical profile of the water where the animals were captured from. By this way we may establish the possible linkage between both sources of information, i.e. if there are correlations between the magnitude and pattern of the biomarkers

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responses and the environmental pollutants bioavailability. In this context biomarkers can act as a nexus between the cause and effect of exposition to a multiplicity of stressful environmental chemicals, thus providing the way to carry out an early overall assessment of these effects at the subindividual level. Responses of effect biomarkers are restricted to the lower levels of biological complexity, i.e. molecular, cellular and tissue levels. Thus, they precede alterations at population of communities’ levels and, consequently may be considered as ‘‘early warning’’ signals in environmental quality assessment. The significant deviation in a biomarker of effect compared to its basal or reference value provide additional information regarding the extent of the impacts caused by stressful environmental agents. In the case of fish, foreign chemicals present in their environment may act as stressors threatening or disturbing their organism; in that case, animal homeostasis initiate an integrated response, i.e. a set of compensating and adaptive responses that readjust metabolic processes in order to cope with the effects of pollutants. Compensating for the effects of chemical stressors causes reallocation of metabolic energy away from investment activities and toward the restoration of homeostasis. The restoration of the damaged homeostasis in a stressed fish is associated with and increase in metabolic rate in relation to the non-stressed condition (Beyers et al., 1999). In addition, research into the recovery of altered biomarkers back to original levels after transfer to clean media provides information in connection with the reversibility of the detected changes if adequate remediation measures are applied. This information is also useful in the prediction of the capacity of a particular population to survive in disturbed environments. The river Reconquista is an urban watercourse in the Province of Buenos Aires (Argentina) approximately 50 km in length. The river basin, traversing what is known as the Pampa region, is sedimentary, flat in topography, with a total surface of 167 000 Ha. For over a century, complex mixtures of domestic, agricultural and industrial solid and liquid wastes (mostly untreated) have been dumped in the river, which has thus become a typical example of the adverse impact of human activities on the health of aquatic environments. Water quality deteriorates along the length of the river, from its source until it flows into the river Luja´n. It should be noted that the river currently has no contamination-free areas (Garcı´a et al., 1998; de la Torre et al., 2002; Olguı´n et al., 2004; Salibia´n, 2006). The test organism used was Cnesterodon decemmaculatus. This is a viviparous species found all year in many places of the river, as well as in many of its effluents in this river basin; the abundance of this species, and its ease of capture were reasons leading to it being chosen for this study. These features ensure the permanent presence of natural populations in environments with different degrees of deterioration, thus providing the means to establish bioindicator parameter baselines for the species using individ-

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uals from less contaminated sites and to compare them with more highly contaminated sites. C. decemmaculatus is a native species which was studied at our laboratory and found suitable as a test organism in acute and chronic aquatic toxicity bioassays (de la Torre et al., 1997, 2002, 2005; Ferrari et al., 1998; Garcı´a et al., 1998). The objective of this study was to assess the toxic impact of the Reconquista river water on a teleost considered representative of local ichthyofauna, considering the early deviations noted in a set of selected biomarkers of effect and their recovery once animals were transferred to unpolluted media as a consequence of an activation of compensatory responses. By screening biomarkers of effect responses, important information will be obtained about organism stress. The measured biomarkers were selected considering their physiological significance as indicators of the basic aspects of the animals’ homeostatic mechanisms. Thus, gill (Na+ + K+)-ATPase activity was considered as an index of the ionoregulatory capacity status, the hepatic key enzymes activities as indicators of changes in the protein metabolism and the brain AChE activity considered as a biomarker target for assessing the exposure to its inhibitors. In addition, a couple of morphological indexes were also checked. The recovery tests were performed comparatively making connections between both the laboratory and field results. 2. Materials and methods 2.1. Animal source and experimental design Female adult C. decemmaculatus were used (total length: 3.0 ± 0.2 cm, body weight: 0.28 ± 0.06 g (means ± SD, n = 69)) . The fish were captured at two different sites: (1) San Francisco: (SF), located 21 km from the source of the river, (2) Artificial pond at the University campus. The community of organisms found in the pond was similar to the one found in nearby natural brooks. Physicochemical characterization of the pond water showed that it is free of contaminants, and was therefore taken as the control site (de la Torre, 2001). Two captures in San Francisco were carried out during the spring (SF1) and late spring (SF2). The fish were assigned into two subgroups: (a) animals used immediately to determine biomarkers. These were placed in glass jars with ice-cold tap-water and transported on ice to the laboratory for processing within a period no longer than 2 h and (b) animals placed in containers with river water and transported to the recovery sites. Two types of recovery tests were performed: (1) Recovery tests in the laboratory. The fish captured in San Francisco (Rtw) were placed in 15-l glass aquariums with a load density no higher than 0.5 g/l; they were kept

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there for 30 days under controlled and constant environmental conditions: (12/12 D/N; 20 ± 1 C), continuous airing with a flow of unchlorinated tap-water and, ad libitum supply of commercially available fish food five times a week. The control group (Ctw) was made up of fish captured in the University campus pond, which were kept under the same conditions after an adaptation period of 15 days. (2) Recovery tests in an outdoor pond. In this case the fish (Rpond) were confined for 41 days inside 5-dm3 perforated plastic boxes, located 0.5 m above the bottom of the campus pond. Load density in these cages did not exceed 1 g/l. Control fish were taken from the same pond Cpond (considered as a reference site). 2.2. Physicochemical parameters Surface water samples were collected from the river and the pond and taken to the laboratory in a period not exceeding 5 h. Samples of the tap-water used in assays were also taken. The samples were kept cool until processed. All the samples were tested for conductivity, alkalinity, hardness, pH, DO, COD, phenols, phosphates, nitrites, ammonium, chlorides, dissolved heavy metals (Cd, Cr, Cu, Pb and Zn) and organochlorine and organophosphorous insecticides. The tests were performed following APHA procedures (1995). Water sampling and fish capture were conducted always simultaneously. 2.3. Water pollution indices The quality of the water samples taken simultaneously with the capture of fish both in the river sampling sites and in the controls (tap-water and campus pond) was determined by means of two local chemical pollution indices, on the basis of relevant selected physicochemical parameters. The ICA (in Spanish: Water Pollution Index) (Bero´n, 1984) is an organic pollution index integrating several physicochemical variables in a weighted algorithm considering all or some of those variables. This index is determined by temperature, Cl, N–NHþ 4 , BOD and DO, and varies between 0 and 10 (worst-best condition); an Index equal to 0 corresponds to a pollution status equivalent to a sewer effluent. The ICAPI (in Spanish: Index of Water Pollution by Industries) (Lacoste and Collasius, 1995) is a numerical expression of contamination due to industrial effluents based on the concentration of phenols, detergents, heavy metals, DO and COD. ICAPI values also vary on a 0–10 scale (worst-best condition), with 0 registering the Index for untreated industrial effluents. In both indices, an Index rate of 10 applies to a water sample in its original state of purity. 2.4. Processing of biological samples Each individual was placed on a cooled glass plate and immediately measured (±0.1 cm) and weighed (±0.01 g).

A dorsal incision was then performed at the opercula, and the gills and brain were removed. The liver was accessed via a ventral incision .The organs were weighed (±0.01 mg). Two morphometric parameters were calculated: the Liver Somatic Index (LSI) and the Condition Factor (CF); the former was determined using the formula LSI = [liver wet weight (g) · total body weight (g)1] · 100 (Sloof et al., 1983) and the latter using the formula CF = body weight (g) · [length (cm)3]1 (Bagenal and Tesch, 1978). The specific activities of the following enzymes were determined: gill ATPase (Na+ + K+) (ATP phosphohydrolase, EC 3.6.1.3), brain aceytylcholinesterase (AChE, acetylcholine acetylhydrolase, EC 3.1.1.7) and liver aminotransferases (AspAT) (L-aspartate: 2-oxoglutarate aminotransferase; EC 2.6.1.1) and (AlaAT) (L-alanine: 2oxoglutarate aminotransferase; EC 2.6.1.2). The following techniques were used: (a) (Na+ + K+)-ATPase: gill filaments were dissected free from the arch, placed into SI buffer (0.3 M sucrose, 0.02 M EDTA, 0.1 M imidazole, pH 7.1) and maintained at 25 C until enzyme activities were measured following a technique based on the Zaugg method (1982). Filaments of the thawed samples were disorganized in a manual glass–glass conical homogenizer maintained on ice. Homogenates were centrifuged at 4 C for 15 min at 10 000g, the supernatant was discarded and the pellets were thoroughly suspended in buffer SI with 0.1% sodium deoxycholate and then centrifuged at 4 C for 15 min at 10 000g. The final supernatants were incubated in the assay medium containing Na+: 130 mM; K+: 20 mM; Mg2+: 4 mM. with and without ouabain (1 mM). The reaction was started by adding ATP (3 mM) and stopped with trichloroacetic acid 30%. The difference in Pi released in both media was considered as the (Na+ + K+) ATPase activity. Hydrolyzed Pi was measured according to Fiske and Subbarow (1925). Results were expressed as lmol Pi mg protein1. (b) AChE: whole brains were homogenized in cold 0.1 M phosphate buffer (PO4HNa2/PO4H2Na), pH 8.0 using a homogenizer maintained on ice with eight strokes at 5000 rpm. The homogenates were centrifuged in a Sorvall RC-5B refrigerated centrifuge at 4 C for 15 min at 10 000g; Pellets were discarded and the enzyme activity was determined on the supernatant according to the Ellman et al. (1961) technique. Results were expressed as nmol min1 mg protein1. (c) Aminotransferases (AspAT, AlaAT): whole livers were homogenized in sucrose 0.25 M (2% w/v). Activities were estimated according to Reitman and Frankel (1957). The reaction mixture contained 2 mmol l1 a-ketoglutarate, AspAT and AlaAT specific substrates (100 and 200 mmol l1 of aspartate

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and alanine l1, respectively) in buffer phosphate (100 mM; pH 7,4). The reaction was started by adding aliquots of the homogenate; after 30 min of incubation, the 2.4-dinitrophenylhydrazine reagent was added an the colored product was evaluated spectrophotometrically at 505 nm. Results were expressed in Karmen units mg protein1. Tissues protein content was determined in duplicate on the final supernatants following the method described by Lowry et al. (1951) using bovine serum albumin as the standard; results were indicated as mg g1 fresh tissue. All measurements were carried out in duplicate. All reagents were analytical grade. 2.5. Statistical analysis Statistical differences between groups were determined using ANOVA (Zar, 1999). The null hypothesis (Ho) was rejected when p < 0.05. Normality and homogeneity of variances were verified using the Kolmogorov–Smirnov test and the Levene test respectively. The Sigmastat 3.1 statistical package was used. 3. Results 3.1. Physicochemical parameters of water samples – pollution status The physicochemical parameters of the river water, tapwater and campus pond water used are set out in Table 1. The physicochemical profile of the river water samples (SF1 and SF2) shows a significant contaminating load, with high levels of COD, nitrites, phenols, and heavy metals (Cr, Cu, Pb and Zn) in addition to a low DO rate. No insecticides were detected in any sample. The phenols, nitrites and heavy metals content exceeded the guidelines water quality levels for freshwater life protection according to Argentine legislation. The results obtained after the estimation of the pollution indices provided additional evidence of the low water quality at the sites where the fish were captured. Both indices showed values corresponding to a polluted – highly polluted level; the ICA ranged between 4.8 and 5.8 while the ICAPI ranged between 3.4 and 3.2. It is worth mentioning that comparable values were previously recorded by Topalia´n and Castan˜e´ (2003) while studying the water quality at other points of the same river, close to San Francisco. Likewise, the ICA and ICAPI values calculated for control samples ranged between 9.5 and 9.8. Physicochemical parameters registered in the pond water (Cpond) were similar to those found in tap-water (Ctw); it should be noted that Zn concentration both in tap-water and pond water was higher than that allowed by the above mentioned law; other measurements carried out in our laboratory confirm these figures (de la Torre et al., 2000; Demichelis et al., 2001; Olguı´n et al., 2004;

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Table 1 Physicochemical profile of Reconquista river water samples taken in San Francisco (SF1 and SF2), in a pond located in the University campus (Cpond) and tap water (Ctw) Parameter

SF1

SF2

Ctw

Cpond

Temperature (C) pH Alkalinity (mM CaCO3) Hardness (mM CaCO3) DO (mg O2 l1) Conductivity (lS cm1) 1 Ammonium (mg N–NHþ 4 l ) 1 Nitrites (mg N–NO 2 l ) Chlorides (mg l1) COD (mg O2 l1) 1 Phosphates (mg PO3 4 l ) Phenols (mg l1)

18.0 8.0 9.2 1.1 2.2 810 s.m. 0.82 50 52 8.1 0.4

21.8 7.9 5.5 0.8 0.6 561 10.6 0.10 33 68 7.4 0.4

20.0 7.5 8.5 0.8 8.2 716 <0.8 <0.02 s.m. <10 <0.1 <0.1

23.2 8.7 7.7 1.1 8.2 753 <0.8 0.04 21 <10 <0.1 <0.1

Heavy metals (lg l ) Cd Cr Cu Pb Zn

2.7 12.7 10.7 7.5 92.8

2.5 4.2 6.1 9.43 165

<2.0 4.0 <2.0 <7.0 66.4

<1 <2 <1 <5 90

Insecticidesb (lg l1) Organophosphates Organochlorines

n.d. n.d.

n.d. n.d.

n.d. n.d.

n.d. n.d.

WPI ICA ICAPI

4.8 3.4

5.8 3.2

9.5 9.8

9.8 9.8

MPQa

1.13 0.06

0.01

1

2.0 2.0 0.8 2.0 30.0

n.d., not detected. WPI, water pollution indices. a MPQ, maximum permitted quantity allowed by the current Argentine legislation for the protection of freshwater aquatic life. b The screening included: (i) Organochlorines: a, b and c HCH, aldrin, endrin, heptachlor epoxide, op 0 and pp 0 DDE, a and c chlordane, endosulfan, op 0 and pp 0 DDT, dieldrin and (ii) organophosphates: ethyl and methyl parathion, fenitrothion, chlorpyrifos (detection limits were 0.1 lg l1).

Castan˜e´ et al., 2006). This could be attributed to the hydrogeological features of the river basin (Pereyra and Tchilinguirian, 2003) added to the anthropic contamination of the aquifers in the basin, associated mainly to industrial activities and intensive agricultural use, which is shown in the significant increase of the concentration of certain heavy metals (Momo et al., 1999; Kruse et al., 2003; Silva Busso and Santa Cruz, 2005). 3.2. Fish biomarkers No mortality was observed throughout the experiments. Tables 2 and 3 show the results of the biochemical and morphological biomarker parameters assessed in fish in each one of the tests. The biochemical tests performed on tissues of fish gathered in the river (SF1; SF2) indicated: (a) Significant inhibition (between 35% and 48%) in gill (Na+ + K+)-ATPase compared to the activity registered in control animals. After transferal to clean media (Rtw; Rpond) the enzyme showed a noticeable

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Table 2 Biochemical parameters of gills, brain and liver, tissue protein content, Liver Somatic Index and Condition Factor of adult females of Cnesterodon decemmaculatus caught in San Francisco and transferred to tap water Parameter Gills (Na+ + K+)ATPase Protein Brain AChE Protein

SF1

Rtw

Ctw

4.7 ± 1.0 (5)

8.7 ± 1.8 (4)

7.2 ± 1.0 (4)

13.5 ± 0.6 (5)

13.5 ± 0.6 (4)

14.0 ± 1.0 (4)

12*

303 ± 11 (10) 39.9 ± 1.8* (10)

313 ± 16 (9) 32.6 ± 1.8 (9)

1576 ± 157* (9) 2338 ± 125 (9)

LSI Protein

1191 ± 98* (8) 3121 ± 174* (8) 0.46 ± 0.08* (8) 2.6 ± 0.2* (8) 188.3 ± 5.9 (8)

2120 ± 96 (9) 2320 ± 164 (9) 0.94 ± 0.06 (9) 1.8 ± 0.2 (9) 178.2 ± 6.5 (9)

Condition Factor

0.95 ± 0.05 (8)

AspAT/AlaAT

SF2

Rpond

Cpond

4.0 ± 1.0* (4)

6.8 ± 0.8 (5)

7.7 ± 0.7 (10)

13.5 ± 0.9 (4)

16.9 ± 0.9 (5)

16.5 ± 0.8 (10)

210.9 ± 10.6* (10) 42.3 ± 1.8 (10)

277.8 ± 14.0 (10) 40.6 ± 3.9 (10)

301.3± 8.5 (18) 37.1 ± 1.2 (18)

AspAT/AlaAT LSI Protein

1938.8 ± 162.9* (9) 3671.0 ± 345.9* (9) 0.55 ± 0.05 (9) 2.1 ± 0.3 (9) 156.8 ± 4.1 (9)

1247.5 ± 49.6* (8) 3656.9 ± 204.1* (9) 0.34 ± 0.03* (8) 1.3 ± 0.1 (9) 152.4 ± 8.0* (9)

1584.7 ± 67.9 (19) 2658.7± 155.9 (19) 0.63 ± 0.04 (19) 1.6 ± 0.2 (19) 180.1 ± 7.5 (19)

Condition Factor

1.16 ± 0.04 (10)

1.11 ± 0.04 (9)

1.11 ± 0.02 (19)

Parameter

227 ± (10) 30.8 ± 2.0 (10)

Liver AspAT AlaAT

Table 3 Biochemical parameters of gills, brain and liver, tissue protein content, Liver Somatic Index and Condition Factor of adult females of Cnesterodon decemmaculatus caught in San Francisco and transferred to an artificial pond

0.68 ± 0.08* (9) 2.5 ± 0.1* (9) 149.2 ± 10.4* (9) 0.88 ± 0.03* (9)

0.93 ± 0.08 (9)

Gills (Na+ + K+)ATPase Protein Brain AChE Protein Liver AspAT AlaAT

Recovery bioassay conducted under laboratory conditions. Ctw, laboratory controls; SF1, fish caught in San Francisco; Rtw, fish recovered in laboratory. Data are means ± SEM; number of samples in parentheses. Gill ATPases in l mol Pi min1 mg protein1 (pool of two individuals); brain AChE in nmol min1 mg protein1, hepatic AspAT and AlaAT in Karmen units mg1 protein; LSI, liver somatic index (%); protein content in mg g1 fresh weight. Significant differences (p < 0.05) between groups are indicated with an asterisk SF1 vs. Ctw; Rtw vs. Ctw.

Recovery bioassay conducted outdoor cages. Cpond, controls kept in artificial pond; SF2, fish caught in San Francisco, Rpond, recovered fish kept in submerged cages in the artificial pond. Data are means ± SEM; number of samples in parentheses. Gill ATPases in l mol Pi min1 mg protein1 (pool of two individuals); brain AChE in nmol min1 mg protein 1, hepatic AspAT and AlaAT in Karmen units mg1 protein; LSI, liver somatic index (%); protein content in mg g1 fresh weight. Significant differences (p < 0.05) between groups are indicated with an asterisk SF2 vs. Cpond; Rtw vs. Cpond.

trend towards increasing its specific activity, reaching the same levels as the corresponding controls (Ctw; Cpond). No differences were noted in the protein tissue content. (b) Important inhibition (between 35% and 38%) in specific brain AChE activity, and a clear tendency towards recovery up to values comparable to those of their controls after transfer to clean media. Protein content in brain tissue remained stable in both experimental series. (c) In both tests aminotransferases were significantly different from those found in the controls. With the exception of AlaAT activity in the laboratory test, basal levels were not reestablished by the end of the recovery period. These discrepancies could be explained by the difference in time spent in clean media. Changes in the AspAT/AlaAT ratio could be interpreted as an indication of the extent of liver damage as control levels were not reached in any of the cases studied.

recovery, however, was not complete. CFs, on the other hand, showed no noteworthy differences in either case.

3.3. Fish biological parameters LSI measured in both groups of fish taken from the river showed significant increases (between 31% and 45%);

4. Discussion On the basis of previous research carried out on the Reconquista river, San Francisco has been characterized as a transition location on the river’s contamination gradient (Castan˜e´ et al., 2006).The values determined for pollution indices support the conclusion that river water quality shows an important deterioration which can be interpreted as a consequence of mixed sewer and industrial pollutants input into the river. Besides, the same indices calculated for control samples (tap-water and campus pond) indicate that water quality there was close to conditions of original purity. Because of their ectothermic condition, the role of temperature in the fish environment may play a critical role; in our case the water temperature range was almost the same both in the field and in the laboratory aquaria. Similarly, water pH and hardness are important factors that determine the bioavailability and toxicity of numerous environmental stressors (Rand et al., 1995). In our study those parameters were similar at all sites. Therefore the differences detected in the samples cannot be attributed to the effect of those parameters.

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Although ionoregulation in fish is mediated by a number of structures including the gastrointestinal epithelium and kidney, ion movement balance mainly takes place in the gills. Gill epithelium is the first interfaces of the organism exposed to the aquatic environment and for this reason a primary and most important target tissue for the action of water-borne pollutants on fish. Gill (Na+ + K+)-ATPase is a membrane-bound enzyme that catalyzes the active Na+ and K+ transport into the animals, providing a driving force in the gill epithelium. The (Na+ + K+)-ATPase activity appeared to be noticeably inhibited by 35–40%, which would suggest the animals most likely suffered disruption, as regards ion regulation. It was estimated that approximately 3% of their resting energy use of fish was spent on Na+, Cl, and osmotic regulation. Hence, even if, as in our case, the secondary reduction in ionoregulatory capacity did not result in death, it is likely that the metabolic costs associated with attempts at ion and water regulation will have considerably increased. Comparable results were found in Cyprinus carpio when exposed to the same river water (de la Torre et al., 1999). Activity for this biomarker showed a clear tendency to return to control values 30–40 days after transferal to clean media, both in laboratory and field and at comparable rates (70% and 85%). The presence of a number of xenobiotics that can adversely affect gill (Na+ + K+)-ATPase activity were detected in the samples of he river water; that was the case of NHþ 4 , nitrites, phosphates, phenols and heavy metals. Waterborne stressors have been reported to increase, to have no effect or to reduce the (Na+ + K+)-ATPase activity during acute or chronic exposure to stressors (Watson and Beamish, 1980; Watson and Benson, 1987; Wendelaar Bonga, 1997). These responses of fish are dependent on the intensity of the stressors’ effects rather than to its type, and may be the consequence of direct effects on the functional and morphological integrity of the epithelial transport (as the disruption of the branchial Cl uptake mechanism provoked by nitrites) or, indirectly, due to the rupture of the endocrine system (for instance, rises in plasma cortisol and catecholamines) (Hanke et al., 1983; Jensen, 2003). AChE is well established as a biomarker of exposure to organophosphorous compounds (OPs) in freshwater fish (Bocquene´ et al., 1997) . The inhibition of acetylcholinesterase in fish brain has been suggested as a variable that might indicate exposure to OPs (Sturm et al., 1999; Dembele´ et al., 2000; de la Torre et al., 2002).The toxic action of OPs on the nervous function is based on the inhibition of the enzyme in brain and muscle. It should be noted that although brain AChE was found to be inhibited by an important percentage, there was nothing to suggest this condition was associated to the existence of organophosphorous insecticides in the water at the location where the fish were captured. Similar findings were reported in previous studies carried out in our laboratory (Rovedatti et al., 2001). Given the fact that contamination in this river

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is characterized by relatively regular discharges (Topalia´n et al., 1999; Rovedatti et al., 2001) and bearing in mind the short half-life of these insecticides in water (Eichelberger and Lichtenberg, 1971), the possibility of these animals having been exposed to these compounds shortly before they were captured should not be dismissed. In this respect, we cannot disregard the inhibitory effect of heavy metals. It is worthwhile to mention that As, Cd and Cu has been shown to depress the AChE in Pimephales promelas (Bocquene´ et al., 1997) at much lower concentration than those measured in our river samples; the decline in AChE activity can be interpreted both as a secondary outcome of conformational changes in the enzyme due to binding metals with functional sulfhydryl groups and as a direct effect derived from their ability to bind covalently to serine at the active AChE site (Szabo and Nemcso´k, 1992). On the contrary, Gill et al. (1991) and Dethloff et al. (1999) reported an activation of brain activity of fish exposed to Cd and Cu. Besides the heavy metals, it is known that other molecules coexisting in the natural environments affected or provoked by anthropogenic activities (e.g. agricultural chemicals like pyrethroids or phytotoxins secondary to algal blooms associated to eutrophication processes) may also be responsible of inhibitory responses of the cholinesterases. It should also be pointed out that inhibition of fish brain AChE can be detected soon after the beginning of exposure to OPs, but the time required for recovery of basal values can take several weeks, being related to the level of depression (Morgan et al., 1990; Sturm et al., 1999). It is interesting that Loteste et al. (2002) found that plasmatic AChE in young Prochilodus lineatus exposed to organophosphorous insecticides showed a similar percentage of inhibition 96 h after being transferred to clean media. Dembe´le´ et al. (1999) studied the recovery pattern of inhibited Cyprinus carpio brain AChE activity and recorded that it took two weeks to return to levels comparable with those registered in controls not exposed to a formulated organophosphate. Comparable results were reported by Ferrari et al. (2004). In our case the activities exhibited a full recovery up to control values after the transference of fish to clean media either in laboratory conditions or in the reference pond. It has been shown that AChE activity may be influenced by the sex and body size of the animals used (Beauvais et al., 2002); since in this study fish of one sex and homogeneous in size were used, these two factors should be disregarded. Similarly, since our study was carried out within a short period of time during spring, our results cannot be attributed to seasonal variations (Moreira and Guilhermino, 2005). More recently, Chuiko et al. (2003) showed important cross-species and cross-families differences in the brain AChE activity of a number of freshwater teleosts. The increase in brain tissue protein content in fish from Rtw group was significantly different from the controls and from the animals captured in the river (SF1). This would seem to point to a compensatory response associated to the synthesis of AChE up to levels close to those of the

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controls. These changes were not detected in the outdoor recovery test. The liver is a central organ of metabolism and is key to the regulation of carbohydrate metabolism. Liver transaminases play an important role in the chemical regulation of intracellular amino acid pools acting as link enzymes between protein and carbohydrate metabolism in all vertebrate classes. The responses to changes in environmental conditions shown by these enzymes could be an index of stress and of metabolic adaptive readjustments. In our case, the assessed parameters showed alterations in the hepatic function of the river fish. Compared to the controls, AlaAT activity was found to be significantly high, 35–38% for SF1 and SF2 respectively. However, complete recovery took place only in animals kept under laboratory conditions. The changes observed in AspAT were significant and in dissimilar way for SF1 and SF2, but in both cases the recovery periods were not sufficient to reach control values. The LSI values in the two control groups were very similar while those of fish captured in the river showed very significant increases. This fact suggest the adverse impact on hepatic function of fish living in a contaminated media that may be interpreted as a consequence of the liver enlargement secondary to exposure to pollutants due to compensatory proliferation processes (Williams and Iatropoulos, 2002). Similar results were reported by Eastwood and Couture (2002) in Perca flavescens captured in environments polluted with heavy metals. This index returned to values close to those recorded for their respective controls only in one case; this result could be attributable to the differences in the time of permanence in clean media. The Condition Factor is accepted as a quantitative indicator of the general physical condition of fish. In our case, this parameter did not display important changes. The results here reported referred to the recovery of the altered parameters resulting from an integrated combination of active monitoring semichronic (caging experiments) and passive monitoring (sampling of wild fish). They did not provide the means to establish a causal relationship between the biomarkers responses and particular substances present in the river water. In fact, the effects should be interpreted as a result of the joint action of specific and unspecific toxic factors. It should not be necessary to link a biochemical response with the presence of a particular contaminant. Because of the interaction of compounds that may exhibit unpredictable toxicities due to their in situ relationships that may either enhance or inhibit their toxicity, it should be recognized that it is not an easy task to determine the specificity of the evaluated responses by means of the used experimental design. However, the range of the detected changes should give an acceptable indication of the severity of the stressors effects present in the analyzed water samples and, consequently, of the intensity of the environmental stress in the particular sampled sites.

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