SciELO - Scientific Electronic Library Online

vol.30 número1Characterization of the Brain-pituitary axis in pejerrey Odontesthes bonariensisSilversides in South Brazil: Morphological and ecological aspects índice de autoresíndice de materiabúsqueda de artículos
Home Pagelista alfabética de revistas  

Servicios Personalizados




  • No hay articulos citadosCitado por SciELO

Links relacionados



versión impresa ISSN 0327-9545

Biocell v.30 n.1 Mendoza ene./abr. 2006


Ecotoxicological studies on the pejerrey (Odontesthes bonariensis, Pisces Atherinopsidae)

Pedro Carriquiriborde and Alicia Ronco

Centro de Investigaciones del Medio Ambiente, Facultad de Ciencias Exactas, UNLP - CONICET. Calle 47 y 115, (1900) La Plata, Bs. As., Argentina.

Address correspondence to: Dr. Pedro Carriquiriborde. Centro de Investigaciones del Medio Ambiente, Facultad de Ciencias Exactas, UNLP, Calle 47 y 115, (1900) La Plata, Bs. As., ARGENTINA. Tel/Fax: (+54-221) 4229329. E-mail:

Key words: Odontesthes bonariensis; Ecotoxicology; Environmental pollution; Heavy metals; Pesticides.

The pejerrey (Odontesthes bonariensis, Cuvier and Valenciennes, 1835) is an Atherinid fish characteristic of the Southern sector of the ‘del Plata' basin (Bonetto and Castello, 1985). This species not only possess ecological relevance (it could represent the largest fish biomass, especially in low diversity euryhaline lakes; Baigún and Delfino, 2001), but also has regional socioeconomic relevance due to commercial and sport fishing, the last representing an annual economic movement of $72 million for the Buenos Aires Province alone (Lopez et al., 2001). In addition, pejerrey is a promissory species for the aquaculture as a result of the high quality of its flesh, appreciated by local and foreign consumers; a fact that in recent years has driven important advances related to the larvae-culture (Miranda and Somoza, 2003), and caged culture (Colautti and Remes Lenicov, 2001) of the species.
On the other hand, pejerrey inhabit rivers and lakes located in the most densely populated and therefore most highly modified regions of Argentina due to human activities. These aquatic systems are usually natural receptors of pollutants released to the environment from point (industry, sewage) and non point (agriculture, urban runoff) sources. Previous studies testify to differing degrees of organic and inorganic contamination of these water bodies and their biota (Ringuelet, 1967; Catoggio, 1990; AGOSBA-OSN-SIHN, 1992; Ronco et al., 1995; Kreimer et al., 1996; Colombo et al., 2000; Menone et al., 2000; Villar et al., 2001).
Despite the ecological and socioeconomical relevance of the pejerrey, and the evidence of differing degrees of pollution in the water bodies that the species inhabit, no published studies on the potential effects of environmental pollutants on the wellbeing of the pejerrey were found at the beginning of our investigations in 1997, and little information on this subject is yet available. As a consequence, the aim of this publication is to review the advances in ecotoxicological aspects of O. bonariensis obtained up to the present in our laboratory (Carriquiriborde, 2004). The studies covered aspects such as lethal and sublethal responses from exposure to some organic and inorganic pollutants of environmental relevance, the influence of some physicochemical characteristics of the water on the toxicity of metals, the bioconcentration pattern of metals according to chemical speciation, and the vulnerability of the species in representative areas of the Southern sector of the del Plata basin.

Methodological Approach

A summary of the ecotoxicological studies conducted during the last seven years by the working group of CIMA to assess adverse effects of pollutants of environmental relevance on O. bonariensis is shown in Table 1. The general conceptual approach for studies included the following aspects: 1) Evaluation of acute lethal effects of the selected toxicants on the species, in order to understand the relative toxicity of each contaminant, and the comparative sensitivity of O. bonariensis in relation to other fish species and aquatic organisms. Additionally, it provided information on the concentration/ response and time/response functions of the selected metals and pesticides. Acute lethal toxicity tests studies were also used to assess the dependence of the adverse effects of metals in relation to the physicochemical characteristics of the water and the source and age of the fish, 2) Evaluation of sublethal effects related to the toxicokinetics and toxicodynamics of the contaminants in order to quantify the bioconcentration of chemicals and their effects at the cellular levels in two main target organs after waterborne exposure to chemicals; the gills and the liver. Effects of chemicals at the cellular levels were investigated through the assessment of the response of the nucleolar activity and the DNA integrity. Selected biological responses were also evaluated as potential biomarkers of exposure and/or effect, 3) Evaluation of the vulnerability of O. bonariensis in the natural environment in order to identify the potential risk of lethal and sublethal exposure to the species in its main natural area of distribution. For this purpose, the toxic exposure ratio (TER) (Harrass, 1996) was calculated on the basis of the information reported in previous studies about the physicochemical characteristics and chemical concentrations in representative water bodies comprising the main distribution area of the species, and the sensitivity levels obtained in our studies.

Ecotoxicological studies conducted with O. bonariensis

Study of Lethal Effects

The study of the acute lethal effects of the metals (Cd+2, Cu+2, Cr+6 and Hg+2) and the pesticides (the herbicide glyphosate and the insecticide cypermethrin)1 indicate that O. bonariensis sensitivity lies in the following order: cypermethrin > Hg+2 > Cd+2 > Cu+2 > Cr+6> glyphosate (Table 2). Sensitivity to cypermethrin was two orders of magnitude greater than Hg(II) and six orders of magnitude greater than glyphosate, the least toxic of the tested chemicals.

Relative sensitivity of O. bonariensis respect to a selected group of chemicals with environmental relevance, and in comparison with other well studied fish species and aquatic organisms.

The LC50 values indicate that no general rule can be made relating to the nature of the chemical (organic or metal) and the relative toxicity on O. bonariensis. The observed relative order of toxicity for the tested chemicals was equivalent to that reported for Oncorhynchus mykiss and also for the mean available LC50 data for fish species. However, it was slightly different to that observed for Cyprinus carpio, which showed an inversion of the sensitivity to Cd and Cu. The lower sensitivity to Cd and higher to Cu was also a pattern reported for crustaceans, and distinctively different from that observed for O. bonariensis (Table 2).
The relative sensitivity of O. bonariensis in comparison with other fish species was, with the exception of glyphosate, always higher than that reported as the average value for all available fish species. In particular, when O. bonariensis was compared with the sensitivity of two well known species it was found to be closer to that reported for O. mykiss (a sensitive species) than for C. carpio (tolerant species). In relation to other Neotropical fish species, the sensitivity of O. bonariensis to all the studied metals ranged between 2.5 and 11 times higher than that reported for the 15 day old Cichlasoma facetum exposed in hard water (Bulus Rossini and Ronco, 2004). In addition, the sensitivity of O. bonariensis to cypermethrin was also higher than that reported for Cichlasoma dimerus (Domitrovic, 2000), for Aequidens portalegrensis to Hg (Hirt and Domitrovic, 1999) and Cd (Hirt and Domitrovic, 2002), and for Cnesterodon decemmaculatus to Cu (Villar et al., 2000).
In comparison with another relevant group in the aquatic environment such as the crustacean, the sensitivity observed for O. bonariensis was one and two orders of magnitude higher for cypermethrin and Cd, respectively and of the same order of magnitude for Cu, Cr, and glyphosate (Table 2). However, in comparison with D. magna, a micro-crustacean broadly used in toxicity bioassays, O. bonariensis presented higher sensitivity only to cypermethrin, a comparable sensitivity to Cd and glyphosate, and a markedly lower sensitivity to Cu and Cr.
The Figure 1 shows an example of the dependence of the LC50 with the time of exposure2 . The dependence of Cr+6 LC50 values with time show a good agreement with the model derived from the critical body residue concept (McCarty and Mackay, 1993), and was consistent with the fact observed for Cu and Ni, that binding of the metals to fish gills predicts acute toxicity better than free-ion activity (Meyer, 1999). In the figure, it is also possible to observe that the age of the organism and water hardness influence the values of the LC50(t) (a slight but significant increase of Cr+6 toxicity was observed when organisms were younger and the water hardness lower). However, the effect induced by these two variables was accurately interpreted by the same time-response model. A similar behavior was also seen for Cd+2 exposed organisms (data not shown). In contrast with the model presented for Cu+2 by Villar et al. (2000), the developed model would provide a mechanistically based description of the relationship between exposure time and the LC50 for the studied metals.

FIGURE 1. Time dependence of LC50 values for O. bonariensis exposed to Cr(VI). Cr 30 H: 30d old organism exposed to Cr(VI) in hard water (255 mg CaCO3/L); Cr 30 S: 30d old organism exposed to Cr(VI) in soft water (30 mg CaCO3/L); Cr 15 H: 15d old organism exposed to Cr(VI) in hard water (255 mg CaCO3/L); black geometric figures: indicate mean LC50 values; bars: 95% confidence limits; curves: adjusted model to the LC50 values; r2: coefficient of determination; horizontal lines: incipient values (LC50(¥)).

The dependence of the estimated LC50 for O. bonariensis with water hardness was more evident for Cd+2 than for Cr+6, and the relationship between the logarithms of these two variables was satisfactorily described by an empirical model proposed for cadmium by USEPA (2001), showing a positive linear correlation (Fig. 2). The figure also shows that O. bonariensis is within the group of sensitive species to Cd detaching from the more tolerant group (regression lines with higher y-intercept). Also the application of a model developed by Welsh et al. (1996) for Pimephales promelas to describe copper toxicity dependence with the water chemistry was used here to explain the behavior of this metal for the case of pejerrey. This model accurately predicted acute lethal effects of Cu on O. bonariensis exposed at different water hardness (data not shown, see Carriquiriborde and Ronco, 2002). Therefore, these models could be used to assess the risk of metal exposure for the species in the field as a function of physicochemical characteristics of the water bodies.

FIGURE 2. Dependence of Cd(II) LC50 for O. bonariensis with water hardness, and its comparison with other fish species (EPA, 2001).
Symbols: LC50 for different fish species (see insert); solid line and equation: linear regression function fitted for O. bonariensis at different hardness values; dotted lines: linear regression function fitted to Cd(II) LC50(96) data published for other species at different hardness values.

Sublethal effects3

Studied sublethal effects in laboratory exposure experiments covered the assessment of the bioconcentration and the response of nucleolar activity to chromium and cadmium; and the genotoxicity of cadmium, copper and chromium in gill and liver tissues.
Bioconcentration depends on the relation between the uptake and elimination rate of chemicals. The study of the bioconcentration4 of the Cd+2 and Cr+6 in the gills and liver of O. bonariensis showed that both metals bioconcentrate in the studied organs, each one following a characteristic pattern (Fig. 3). Small concentrations of waterborne Cd (» 0.1 mM) were highly accumulated in the gill (1000 fold), but only little Cd was accumulated in the liver (60 fold). On the other hand, relatively high concentrations of waterborne Cr (»20 mM) were moderately bioconcentrated in the gill (only 2.5 fold), but presented low accumulation in the liver.

FIGURE 3. Chromium and cadmium bioconcentration patterns for O. bonariensis after 384h exposure.
Values: represent approximate micromolar concentrations; shades: symbolize concentration levels.

Under the assayed conditions, cadmium is mainly present as the free cation, Cd+2, that quickly binds to gill surface to be later slowly up-taken (Pagenkopf, 1983; Playle 1998) passively through Ca+2 channels (Verbost et al., 1989; Wicklund Glynn et al., 1994). In a different way, chromium is mainly present as the anion CrO4-2/Cr2O7-2 (Kotás and Stasicka, 2000) and it is believed that the metal is up-taken through transporters for SO4 -2 and PO4 -3 (Wetterhahn, 1979; Ottenwälder and Wiegand, 1988). As a consequence, these properties and mechanisms characteristic of each metal could be responsible for the differences observed in the pattern and the magnitude of the bioconcentration of Cd+2 and Cr+6 in the gills and liver of O. bonariensis. The higher affinity of the Cd+2 for the gill surface described by other authors agrees with the higher accumulation of the metal observed in this tissue. Furthermore, this could explain the observed correlation between the model derived from the CBR concept and the LC50(t), and may also explain the higher lethal acute toxicity of Cd+2 in comparison with Cr+6. Moreover, according to Norey et al. (1990), the bioconcentration pattern of Cd observed in O. bonariensis, higher in the gill than in the liver, places the pejerrey within the group of sensitive species such as the salmonids to this metal.
The nucleolar activity is mainly given by the product of ribosome biogenesis, which is directly linked with the protein synthesis of the cell. In the yeast, 75% of the most highly transcribed genes are ribosomal protein genes, representing an important energetic cost for the cell (Jelinsky and Samson, 1999). Consequently, it is tightly regulated (Leary and Huang, 2001; Jorgensen et al., 2002) and is very sensitive to factors that induce cellular stress and DNA damage (Rubbi and Milner, 2003). Our experiments demonstrate that the nucleolar activity5 in the gill and liver of O. bonariensis was affected by sublethal concentration of Cr+6 and Cd+2, also showing a characteristic response depending on the metal and the organ. Whereas a general reduction of the nucleolar activity (assessed by the mean nucleolar area) was induced by Cr+6 both in the gills and the liver, a decrease of this parameter was induced by Cd+2 only in the gill. Conversely, a significant increase of the nucleolar activity was caused by this metal in the liver. The nucleolar activity assessment end points NOEC and LOEC found for each metal and tissue are shown in Table 3.

Summary of sublethal effects induced by Cd, Cr, and Cu in O. bonariensis

DNA integrity6 , assessed by means of the comet assay identifies double and simple strand breaks and labile alkaline sites induced in the DNA by chemicals. Results showed that all the tested metals (Cd+2, Cr+6, Cu+2) were able to induce genotoxic effects on the studied tissues of O. bonariensis. It was evidenced as a significant increase in the frequency of DNA damaged nucleus presented in the gill and liver of metal exposed fish with respect to those in the control group. As was expected from a waterborne exposure, the three metals induced higher DNA damage in the gill than in the liver in a time dependent manner (Carriquiriborde et al., 2002a, 2003). The NOEC and LOEC found for each metal and tissue is shown in Table 3.
The observed sublethal effects can also be linked to the bioconcentration pattern of the Cd and Cr, through the CBR concept. In the gill, organ with the highest concentration of both metals, the nucleolar activity was significantly reduced and was the only organ where DNA damage was observed. Reduction of the nucleolar activity can be related to the demonstrated capacity of Cd+2 at cytotoxic concentrations to induce "in vitro" and "in vivo" general inhibition of cellular proliferation, gene expression, and protein synthesis (Beyersmann and Hechtenberg, 1997). In addition, it was observed that Cd induced DNA damage only in the gill, a reported effect only at cytotoxic concentrations of this metal (Beyersmann and Hechtenberg, 1997). The reduction caused by Cr+6 can be linked to the capability of this metal to induce cross links with proteins and DNA during its intracellular reduction from Cr+6 to Cr+3 (Stearns and Wettterhahn, 1994). In addition, both metals are able to generate general cellular injury by means of oxidative stress (Stohs and Bagchi, 1995). On the other hand, in the liver, while chromium (the higher concentrated of both metals, Fig. 3) still decreased the nucleolar activity and induced DNA damage, cadmium only increased nucleolar activity. The observed increase of the nucleolar activity induced by cadmium in the liver is consistent with the observation that small quantities of the metal induce protein synthesis, gene expression and cell proliferation (Beyersmann and Hechtenberg, 1997). The same authors state that the stimulation induced by cadmium on cellular signals at various stages of mitogenic cascades, protooncogene expression, DNA synthesis, and cell proliferation may be a clue for the interpretation of the carcinogenic action of this metal.

Vulnerability of O. bonariensis in the natural environment

Vulnerability of the species to be adversely affected by the Cd, Cu, Cr, and cypermethrin was assessed considering the information generated on the lethal and sublethal effects of these metals and information on water physical-chemistry and metal concentration reported for water bodies located within the main distribution area of the species.
It was shown in the previous sections that physicochemical characteristics of the water affect metal toxicity on O. bonariensis. The effect of the hardness, pH and DOC on Cd+2 and Cu+2 lethal acute toxicity can be easily assessed by means of empirical models, but no model is yet available for Cr. Consequently, in order to predict more accurately the vulnerability of O. bonariensis to these metals within the main area of distribution of the species, values of hardness, pH, DOC, and metal concentration were collected from the available literature and analyzed.
The reported pH range for the Lower Paraná River was 6.1-7.9 (Villar et al., 1999), for the Río de la Plata 5.8-9.6 (AA-AGOSBA-ILPLA-SHN, 1997), for Samborombón River 8.3-8.4 (Mercado, 2000), for Chascomús Lake 8.0-9.0 (Conzonno and Claverie, 1990), and for San Miguel del Monte pond 8.23-9.20 (Gabellone et al., 2001). In addition, mean hardness values, estimated as the sum of Ca+2 and Mg+2 concentrations reported in the literature, were expressed as calcium carbonate. Hardness interval values found for the Lower Paraná River were 22.5-32.0 mg CaCO3/L (estimated from Villar et al., 1999), and for the Río de la Plata, 22.5-217.5 mg CaCO3/L (estimated from AAAGOSBA- ILPLA-SHN, 1997). Mean hardness values calculated for the lower, middle and upper sectors of the Samborombón River were 86, 87, 155 mg CaCO3/ L (estimated from Mercado, 2000). Hardness ranges obtained for Vitel and Chascomús ponds were 94.9- 205.4 mgCaCO3/L (estimated from Olivier, 1961), and 61.5-161.3 mgCaCO3/L (estimated from Conzonno and Claverie, 1990), respectively. Hardness reported for Salado River (Guerrero's Bridge) was 262 mg CaCO3/ L. Finally, the mean and minimum DOC values reported were 2.8 - 4.2 for the Lower Paraná River and 4.0-7.9 mgC/L for the Río de la Plata (Villar et al., 1999). Minimum and maximum values found for the same parameter in some representative water bodies from the Salado River Basin were: Vitel pond 12.0- 57.0 (Olivier, 1961) and Chascomús pond 7.7-18.7 (Conzonno and Claverie, 1990).
The maximum concentration of Cd, reported for the lower Paraná River, was 1 mg/L (Gómez et al., 1998). For the Río de la Plata, the range of maximum values reported for this metal was 1.0-3.7 mg Cd/L (AGOSBAOSN- SIHN, 1992; Villar et al., 1998), with the highest value found in front of Berazategui. For the Salado River Basin, the observed range in Cd concentration was from 1.0-5.0 mg Cd/L, with higher concentrations found in the upper sector of the basin (Ronco and Argemí, 1998; Carriquiriborde et al., 2002b). In addition, the range of maximum values of Cr reported for the Lower Paraná River and Río de la Plata were 19.7 - 44.2 and 5 - 4,100 mg Cr/L, respectively (Villar et al., 1998; AGOSBAOSN- SIHN, 1992; AA-AGOSBA-ILPLA-SHN, 1997). Extremely high values were observed in the Riachuelo River mouth with concentrations of 72,200 mg Cr/L (Kreimer et al., 1996). The values reported for the Salado River Basin ranged between 35 and 71 mg Cr/L, and again, the higher levels were found in the upper sector of the basin (Ronco and Argemí, 1998; Carriquiriborde et al., 2002b). Moreover, the range of copper concentrations reported within the distribution area of the species were 3-8 mg/L in the lower Paraná River, 4-12 mg/L in Río de la Plata (Villar et al., 1999), and between 12 to 16 in the Salado River Basin, with the highest value observed in del Monte pond (Carriquiriborde et al., 2002b). Finally, the interval of cypermethrin concentrations reported for streams and the mouths of affluent rivers of the lower Paraná River was 1 - 190 μg /L (Marino et al., 2004).
A summary of the physicochemical parameter values corresponding to the worst case scenario (lowest values of hardness, pH, and DOC, and highest metal concentrations) was constructed from the published field data of water bodies and placed within the main distribution area of the species. The LC50(¥) values corrected as a function of these physicochemical parameters and the toxic effect ratio (TER) values according to Harrass (1996), represented as a quotient between the adjusted lethal acute toxicity value and the reported field metal concentration, is displayed in Table 4.

Vulnerability of O. bonariensis to heavy metal acute exposure in the southern sector of the del Plata basin.

According with the data presented above it is possible to assume that: a) the physicochemical characteristics reported for the water bodies of the southern sector of the del Plata basin showed a decrease of hardness and pH values from the South to the North and from the West to the East, showing the lower Paraná as the place of maximum vulnerability for the species to metal exposure, and b) the concentration of metal and cypermethrin near or greater than those that induced biological effects in O. bonariensis were reported for metals in the Río de la Plata and upper Salado River Basin, and for cypermethrin in the lower Paraná River.
The TER, shows the Río de la Plata as the place where risk of lethal acute exposures to the three studied metals for O. bonariensis is the highest, not only as a consequence of the elevated metal level contents reported for its waters, but also resulting from the relatively low values of hardness and pH, and DOC that characterize this water body. In addition, risk of O. bonariensis to Cd and Cu lethal acute exposure should be considered in the lower Paraná River, as well as to Cd in the Salado River Basin. From another point of view, a high risk of acute lethal effects (TER<10) was found for Cd in the whole distribution area, for Cu in Río de la Plata and the Lower Paraná, and for Cr only in Río de la Plata. On the other hand, the lowest TER was obtained for cypermethrin in the mouth of affluent rivers of the lower Paraná River which drain from ‘Pampa Ondulada', one of the most intensely cultivated areas of Argentina, indicating a high risk of exposure for O. bonariensis to this compound related to agricultural activities.

Concluding Remarks and Perspectives

In the present paper, ecotoxicological studies obtained during the last seven years in our laboratory on O. bonariensis, were reviewed. On the basis of the current knowledge, it is possible to state that:
O. bonariensis is a very sensitive species to environmental pollution, and also presents a relative response pattern to chemicals comparable to most fish species.
Acute lethal effects of waterborne Cd+2 and Cr+6 (and probably other metals) on O. bonariensis is mainly related to the ability of each metal to bind to the gill and accumulate there. The magnitude of the metal accumulation in the gills not only depends on its chemical nature but also on its bioavailability, which is conditioned by physicochemical characteristics of the water such as its hardness. The lethal acute toxicity of these metals also markedly depends on the age of the organism and the time of exposure up to 48h but, after this time variation of lethality is much lower.
Sublethal effects of Cd(II) and Cr(VI) (and probably other metals) are also directly related to the distribution and concentration of the metal in the specific tissues. Reduction of the nucleolar activity induced by both metals is in good accordance with the general cellular damage proposed for these metals at cytotoxic concentrations, while the increase of the nucleolar activity induced by Cd may be related with its ability to stimulate protein synthesis, gene expression and cell proliferation at very low concentrations. As a consequence, nucleolar activity seems to be a good indicator of cell arrest or proliferation induced by metals in O. bonariensis. Genotoxic effects of Cd and Cr assessed by the COMET assay are in good correspondence with the general response observed for the nucleolar activity, and with the capacity to induce DNA damage as previously reported for these metals.
• According to the physicochemical characteristics and chemical concentrations reported for the water bodies included in the natural area of distribution of O. bonariensis, and the high sensitivity of the species to the studied environmental pollutants, it is possible to affirm that vulnerability to metal exposure is at a maximum in the soft and slightly acidic waters of the lower Paraná River, and that potential risk of exposure to this class of chemicals exist near urban areas, such as the occidental margin of the Río de la Plata. On the other hand, high potential risk of exposure to cypermethrin exists in rivers draining into agricultura

Although unprecedented information on ecotoxicological aspects of O. bonariensis was summarized in the present paper, it only represents the beginning in understanding the effects of pollutants on this important natural resource and emblematic fish of the shallow ponds of the ‘pampas' region. It is clear that a great amount and diversity of studies will still be necessary to have a clear insight on the impacts of toxicants on the species in its natural environment.
From the results obtained in the reviewed studies it is possible to establish a number of futures lines of research, however, many others would still be relevant. For example, the presented methodology could be used to complete the scheme presented in table 1, or to asses the adverse effects of other pesticides and chemicals of environmental concern. In addition, new experiments could be conducted to asses the more recently proposed Biotic Ligand Model as a tool to evaluate the effect of the physicochemical characteristics on the toxicity of metal. Furthermore, the observed response of the nucleolar activity to metals opens a new field to study possible mechanisms involved in such responses. On the other hand, it would be interesting to conduct field studies in selected regions to assess the potential risk areas and possible effects on O. bonariensis natural populations.
According to current challenges in ecotoxicology, studies with O. bonariensis could also be oriented to assess: i) chronic effects of pollutants, such as endocrine disruption, DNA damage/mutagenesis, deficiencies in the immune system, and neurological effects; ii) multiple effects by single pollutants, as for example multiple target sites and multiple modes of toxic action, and time- and organ-dependent effects; iii) effects of complex mixtures of pollutants, assessing the effects of wastewater treatment plant effluents, field runoff, pollutants and their degradation products, and complexes of chemical compounds; iv) multiple stressors, as for example assessing the join effect of UV, temperature and pathogens on pollutants toxicity; and v) ecosystem complexity, such as variations in species sensitivities, effects of propagation from organisms to populations and ecosystems, and identification of the stressor-effect relationship.
The challenges in ecotoxicology also call for novel bioanalytical methodologies such as modern molecular and genetic tools or sophisticated test systems for routine screening. However, it will be necessary to understand how subtle changes at the molecular level or laboratory scale affect O. bonariensis in its environment. It is evident that further studies will have to be undertaken to provide better tools in the species management and protection.


Financial support for studies was from the National Research Council of Argentina (CONICET) and the University of La Plata. Authors wish to acknowledge F Dulout (one of the PhD Thesis Directors), JC Deluca, S Picco, MP Heras from CIGEBA-UNLP; D Colautti, G Berasain from the Fisheries Office; and F Argemi, JP Streitenberger from CIMA-UNLP for their specific collaboration in different aspects of the studies summarized in this review. Also authors wish to thank the Fisheries Office of the Ministry of Production of the Buenos Aires Province for supplying test organisms, and Mr Roberto Cantarini for the supply of fish food. Cytological studies were performed in the Centre for Basic and Applied Studies on Genetics (CIGEBA), Faculty of Veterinary, UNLP.


1. Lethal acute toxicity tests were conducted following recommendations of the USEPA (1993). Static-renewal tests were employed to evaluate the 50% lethal concentration (LC50) for Cd+2, Cr+6, Cu+2, Hg+2, cypermethrin, and glyphosate. Fifteen or 30-day old organisms were used as test organisms. Assay were performed in hard water, (La Plata tap water; 255.3 mg CaCO3/L) and reconstituted soft water (31.4 mg CaCO3/L). Two hundred and fifty ml polyethylene or glass test chambers filled with 200ml of test solution were utilized for acute lethal toxicity assessment of metals and pesticide, respectively. Toxicity tests were conducted by triplicate using 5 or 10 organisms per replica in preliminary or final tests, respectively. Temperature was kept at 22 ± 1ºC and and pH 7.2. Larvae were fed with 24h Artemia sp. nauplii once at 48h exposure. Dead fish were removed and recorded every 24h during 96h. LC50 were estimated at each recording time by means of the Probit method (Finney, 1971) using a specific software (Probit; USEPA version 1.5).
2. Time dependency of the LC50 values was explained by the following model: LC50(t) = LC50(¥) / [1- e (-ke ¥ t)]2, derived from the Critical Body Residue (CBR) concept and more specifically the Lethal Body Burden (LBB) approach (McCarty, 1986; McCarty and Mackay, 1993).
Effects of physicochemical characteristics of the water on Cd+2 and Cu+2 toxicity were assessed by means of empirical equations obtained from USEPA (2001) and Welsh et al. (1996), respectively. The equation used to explain the relationship between water hardness and Cd+2 LC50 was: Ln(LC50Cd)= 1.205[Ln(hardness)]-3.949. The equation used to explain the relationship between pH, DOC, and hardness and Cu+2 LC50 was: Log(LC50Cu)= 0.192(pH) + 0.136(pH ¥ log DOC) + 0.166(Ca) -0.981, replacing Ca by the hardness, expressed as mg CaCO3/L, as the product of 0.4 ¥ [Ca], then the equation is Log(LC50Cu)= 0.192(pH) + 0.136(pH ¥ log DOC) + 0.0664(Hardness) -0.981. The parameters of these equations were recalculated for O. bonariensis using the LC50 values obtained in the laboratory at different hardness values.

3. Sublethal exposure experiments were performed using six-month old juveniles exposed in 20 L polyethylene test chambers to three sublethal concentrations of Cd+2, Cr+6 or Cu+2 for a maximum period of 384 h. Assays were performed by triplicate using 2 organisms per replica per treatment (time/concentration). Fish were sampled at 6, 48, 192 and 384 h. Static renewal experiments were conducted using La Plata tap water, temperature was kept at 22 ± 1ºC, pH was 7.2, and juveniles were fed with adults of Daphnia sp. every 48h.
4. Bioconcentration of Cd and Cr in gill and liver of O. bonariensis juveniles sampled at 48 and 384h were measured by means of electrothermic absorption spectroscopy using a GBC Scientific Equipment PTY. LTD. Model: AA902 equipped with a graphite furnace atomizer GF2000 and an autosampler, and following the operative manual guidelines. Tissues were previously digested following the method proposed by Handy and Depledge (1999).
5. Nucleolar activity was evaluated in gill and liver cells sampled from juveniles exposed to Cd+2 and Cr+6 during 6, 48, 192, and 384h. Histological preparations and nucleolar area and number assessment were conducted following the method proposed by Arkhipchuk and Palamarchuk (1997) and adapted by Carriquiriborde et al. (2000).
6. DNA integrity in the gill and the liver cells of O. bonariensis juveniles exposed to Cd+2, Cr+6 and Cu+2 for 48 and 348h was assessed using the single cell gel electrophoresis assay (COMET assay) according to Singh et al. (1988). Tissue samples were pretreated following the methodology proposed by Deventer (1996) but using only trypsin to isolate cell from tissue. Qualitative analysis of DNA damage was performed scoring the observed nuclei in 5 levels, 0, I, II, III, IV according to the degree of DNA migration (tail).


AA-AGOSBA-ILPLA-SHN (1997). Calidad de las Aguas de la Franja Costera Sur del Río de la Plata (San Fernando- Magdalena). Consejo Permanente para el Monitoreo de la Calidad de las Aguas de la Franja Costera Sur del Río de la Plata, Bs. As. pp 157.         [ Links ]
AA-AGOSBA-OSN-SHN (1992). Río de La Plata. Calidad de las aguas de la franja costera sur (San Isidro - Magdalena). Servicio de Hidrografía Naval de la Armada Argentina, Bs. As. pp 53.         [ Links ]
Arkhipchuk VV, Palamarchuk JV (1997). Heteromorphic-paired nucleoli in fish cells. Cytobios 88: 229-238.         [ Links ]
Baigún C, Delfino R (2001). Consideraciones y criterios para la evaluación y manejo de pesquerías de pejerrey en lagunas pampásicas. In: Fundamentos biológicos, económicos y sociales para una correcta gestión del recurso pejerrey. F Grossman (ed.). Asyanax, Azul, Bs As, pp. 132-145.         [ Links ]
Beyersmann D, Hechtenberg S (1997). Cadmium, gene regulation, and cellular signaling in mammalian cells. Toxicol Appl Pharmacol 144: 247-261.         [ Links ]
Bonetto AA, Castello HP (1985). Pesca y piscicultura en aguas continentales de América Latina. Progr Des Cient y Téc OEA (Washington), Ser. Biol. Monografía N°31: 1-114.         [ Links ]
Bulus Rossini GD, Ronco AE (2004). Sensitivity of Cichlasoma facetum (Cichlidae, Pisces) to Metals. Bull Environ Contam Toxicol, 72: 763-768.         [ Links ]
Carriquiriborde P, Dulout FN, Ronco AE (2000). Effect of Cr(VI) on nucleolar activity in liver cells of Odontesthes bonariensis (Pisces). Appl Biol Sci 6: 57-65.         [ Links ]
Carriquiriborde P, Ronco A (2002). Sensitivity of the Neotropical Teleost Odontesthes bonariensis (Pisces, Atherinidae) to Chromium(VI), Copper(II), and Cadmium(II). Bull Environ Contam Toxicol. 69: 294-301.         [ Links ]
Carriquiriborde P, Picco S, Ronco AE, De Luca JC (2002a). Evaluación del efecto genotóxico del Cu(II) en hígado y branquia de Odontesthes bonariensis (Pisces; Aterinidae) por medio del ensayo COMETA. XXXI Congreso Argentino de Genética. La Plata, 17-20.         [ Links ]
Carriquiriborde P, Streitenberger JP, Argemi JCF, Ronco AE (2002b). Efecto de contaminantes Ambientales en Pejerrey (Odontesthes bonariensis). Segundas Jornadas Sobre "Ecología y Manejo de Ecosistemas Pampeanos". Sección Biología de Poblaciones, Instituto "Dr Raúl A. Ringuelet, UNLP-CONICET. La Plata, 13-15.         [ Links ]
Carriquiriborde P, Picco SJ, Ronco AE, De Luca JC, Dulout F (2003). Evaluación del efecto genotóxico del Cd(II) en hígado y branquia de Odontesthes bonariensis (Pisces; Atherinidae) por medio del ensayo COMETA. VI Congreso SETAC Latinoamérica. Buenos Aires, Argentina, 20-23.         [ Links ]
Carriquiriborde P (2004). Respuestas Biológicas a Contaminantes Ambientales en Peces: Estudio de la Bioconcentración, de los Efectos Letales y Subletales y la Evaluación de Biomarcadores en Odontesthes bonariensis Expuesto a Diferentes Concentraciones de Cadmio y Cromo en el Agua. Tesis Doctoral, Facultad de Ciencias Naturales y Museo, Universidad Nacional de la Plata.         [ Links ]
Catoggio JA (1990). Contaminación del Agua. Causas de la Contaminación de Aguas Superficiales y Subterraneas. Precipitaciones Acidas. Eutroficación; Polución Costera. In: Latinomerica Medio Ambiente y Desarrollo. Eds. Fundación J. E. Roulet, Buenos Aires. pp. 137-155.         [ Links ]
Colautti D, Remes Lenicov M (2001). Primeros resultados sobre cría de pejerreyes en jaulas: crecimiento, supervivencia, producción y alimentación. In: Fundamentos Biologicos, Economicos y Sociales para una Correcta Gestion del Recurso Pejerrey. Grossman F, Eds., Editorial Astyanax. Azul, AR, pp. 54-64.         [ Links ]
Colombo JC, Bilos C, Remes Lenicov M, Colautti D, Landoni P, Brochu C (2000). Detritivorous fish contamination in the Río de la Plata estuary: a critical accumulation pathway in the cycle of anthropogenic compounds. Can J Fish Aquat Sci 57: 1139-1150.         [ Links ]
Conzonno VH, Claverie EF (1990). Chamical characteristics of the water of Chascomús pond (Provincia de Buenos Aires, Argentina). Limnological implications. Rev Brasil Biol 50: 15-21.         [ Links ]
Deventer K (1996). Detection of genotoxic effects on cell of liver and gill of B. rerio by means of single cell gel electrophoresis. Bull Environ Contam Toxicol 56: 911-918.         [ Links ]
Domitrovic HA (2000). Toxicidad y respuesta histopatológica en Cichlasoma dimerus (Pisces, Cichlidae) expuestos a cipermetrina en ensayos de toxicidad aguda. Comunicaciones Científicas y Tecnológicas 2000. Universidad Nacional del Nordeste. pp 4. v_pdf/v_047.pdf.         [ Links ]
Finney (1971). Probit Analysis, 3rd ed. Cambridge University Press, London.         [ Links ]
Gabellone NA, Solari LC, Claps MC (2001). Planktonic and physico-chemical dynamics of a markedly fluctuating backwater pond associated with a lowland river (Salado river, Buenos Aires, Argentina). Lakes & Reservoirs: research and management 6: 133-142.         [ Links ]
Gómez S, Villar C, Bonetto C (1998). Zinc toxicity in the fish Cnesterodon decemmaculatus in the Paraná River and Río de La Plata Estuary. Environ Poll, 99: 159-165.         [ Links ]
Handy RD, Depledge MH (1999). Physiological responses: their measurements and use as environmental biomarkers in ecotoxicology. Ecotoxicology 8: 329-349.         [ Links ]
Harrass MC (1996). Regulatory use of ecotoxicology statistics: a US perspective. Ecotoxicology 5: 145-154.         [ Links ]
Hirt LM, Domitrovic HA (1999). Toxicidad y Respuesta Histopatológica en Aequidens portalegrensis (Pisces, Cichlidae) Expuestos a Bicloruro de Mercurio en Ensayos de Toxicidad Aguda y Subletales. Comunicaciones Científicas y Tecnológicas 1999. Universidad Nacional del Nordeste. pp 4.
Hirt LM,         [ Links ] Domitrovic HA (2002). Toxicidad y Respuesta Histopatológica en Aequidens portalegrensis (Pisces, Cichlidae) Expuestos a Cloruro de Cadmio en Ensayos de Toxicidad Aguda y Subletales. Comunicaciones Científicas y Tecnológicas 2002. Universidad Nacional del Nordeste. pp 4.
Jelinsky SA,         [ Links ] Samson LD (1999). Global response of Saccharomyces cerevisiae to an alkylating agent. Proc Natl Acad Sci USA 96: 1486-1491.         [ Links ]
Jorgensen P, Nishikawa JL, Breitkreutz BJ, Tyers M (2002). Systematic identification of pathways that couple cell growth and division in yeast. Science 297: 395-400.         [ Links ]
Kotas J, Stasicka Z (2000). Chromium occurrence in the environment and methods of its speciation. Environ Poll 107: 263- 283.         [ Links ]
Kreimer ED, Palacios DE, Ronco AE (1996). A proposal for dredging contaminated sediments at the Dock Sud Port, Argentina. Proceedings of the International Symposium on Coastal Ocean Space Utilization, COSU'96, Buenos Aires, pp. 435-445.         [ Links ]
Leary DJ, Huang S (2001). Regulation of ribosome biogenesis within the nucleolus. FEBS Letters 509: 145-150.         [ Links ]
López H, Baigún C, Iwaszkiw J, Delfino R, Padin O (2001). La Cuenca del Salado: Uso y posibilidades de sus recursos pesqueros. Ed. Universidad de La Plata, La Plata, Bs. As. pp. 60.         [ Links ]
Marino D, Cremonte C, Berkovic A, Ronco A (2004). Análisis de Cipermetrina y Clorpirifos en Aguas, Sedimentos y Suelos de Zonas Agrícolas. Anales XXV Congreso Argentino de Química S8: 1385-1388.         [ Links ]
McCarty LS (1986). The relationship between aquatic toxicity QSARs and bioconcentration for some organics chemicals. Environ Toxicol Chem 5: 1071-1080.         [ Links ]
McCarty LS, Mackay D (1993). Enhancing ecotoxicological modeling and assessment: body residues and modes of toxic action. Environ Sci Technol 27: 1719-1728.         [ Links ]
Menone ML, De Moreno JEA, Moreno VJ, Lanfranchi AL, Metcalfe TL, Metcalfe CD (2000). PCBs, and Organochlorines in tissues of Silverside (Odontesthes bonariensis) from a Coastal Lagoon in Argentina. Arch Environ Contam Toxicol 38: 202-208.         [ Links ]
Mercado L (2000). Evaluación de la calidad de aguas de seis sistemas lóticos pampásicos mediante el estudio de variables físicas t químicas. Rev Mus Argentino Cienc Nat 2: 27-35.         [ Links ]
Meyer JS (1999). A mechanistic explanation for the Ln(LC50) vs. Ln(hardness) adjustment equation for metals. Environ Sci Technol 33: 908-912.         [ Links ]
Miranda LA, Somoza GM (2003). Control de la reproducción del pejerrey Odontesthes bonariensis en cautiverio. V Jornadas Nacionales de Ciencias del Mar. Mar del Plata, Argentina.         [ Links ]
Norey CG, Brown MW, Cryer A, Kay J (1990). A comparison of the accumulation, tissue distribution and secretion of cadmium in different freshwater fish. Comp Biochem Physiol 97: 215- 220.         [ Links ]
Olivier SR (1961). Estudios limnológicos de la laguna de Vitel (partido de Chascomús, Buenos Aires - Argentina). Agro 3: 1-128.         [ Links ]
Orme S, Kegley S (2004). PAN Pesticide Database, Pesticide Action Network, North America (San Francisco, CA. 2004), URL:
Ottenwä         [ Links ]lder H, Wiegand HJ (1988). Uptake of 51Cr(VI) by human erythrocytes: Evidence for a carrier-mediated transport mechanism. Sci Total Environ 71: 561-566.         [ Links ]
Pagenkopf GK (1983). Gill surface interaction model for trace metals toxicity to fish: role of complexation, pH, and water hardness. Environ Sci Technol 17: 342-347.         [ Links ]
Playle RC (1998). Modelling metal interactions at fish gills. Sci Tot Environ 219: 147-163.         [ Links ]
Ringuelet RA (1967). Contaminación o polución del ambiente acuático con referencia especial a la que afecta el área platense. Agro 9: 5-33.         [ Links ]
Ronco AE, Argemí F (1998). Metales pesados en agua, plancton, peces y sedimentos de la Laguna de Lobos, Provincia de Buenos Aires. Gerencia Ambiental 5: 788-792.         [ Links ]
Ronco AE, Sobrero MC, Alzuet PR, Bulus Rossini G, Dutka B (1995). Screening for sediment toxicity in the Río Santiago basin: A base line study. Environ Toxicol Water Qual. 10: 35-39.         [ Links ]
Rubbi CP, Milner J (2003). Disruption of the nucleolus mediates stabilization of p53 in response to DNA damage and other stresses EMBO 22: 6068-6077.         [ Links ]
Singh NP, McCoy MT, Tice RR, Schneider EL (1988). A simple technique for quantitation of low levels of DNA damage in individual cells. Exp Cell Res 175: 123-130.         [ Links ]
Stearns DM, Wetterhahn KE (1994). Reaction of chromium(VI) with ascorbate produces chromium(V), chromium(IV), and carbon-based radicals. Chem Res Toxicol 7: 219-230.         [ Links ]
Stephenson RR (1982). Aquatic Toxicology of Cypermethrin. I. Acute Toxicity to Some Freshwater Fish and Invertebrates in Laboratory Tests. Aquat Toxicol 2: 175-185.         [ Links ]
Stohs SJ, Bagchi D (1995). Oxidative mechanisms in the toxicity of metal ions. Free Radicals Biol Med 18: 321-336.         [ Links ]
USEPA (1993). Methods for Measuring the Acute Toxicity of Effluent and Receiving Water to Freshwater and Marine Organism, 4th ed. Environmental Protection Agency, Washington DC, EPA/600/4-90/027F.         [ Links ]
USEPA (2001). 2001 Update of Ambient Water Quality Criteria for Cadmium. Environmental Protection Agency, Washington DC, EPA-822-R-01-001.         [ Links ]
USEPA (2003). 2003 Draft Update of Ambient Water Quality Criteria for Cadmium. Environmental Protection Agency, Washington DC, EPA 822-R-03-026.         [ Links ]
USEPA (2004). ECOTOX: Ecotoxicology Database, USEPA/ORD/ NHEERL, Mid-Continent Ecology Division. URL:
Verbost PM,         [ Links ] Van Rooij J, Flik G, Lock RAC, Wendelaar Bonga SE (1989). The movement of cadmium through freshwater trout branchial epithelium and its interference with calcium transport. J Exp Biol 145: 185-197.         [ Links ]
Villar C, Stripeikis J, Tudinio M, D'Huicque L, Troccoli O, Bonetto C (1999). Trace Metal Concentrations in Costal marshes of the Lower Paraná River, and the Río de la Plata Estuary. Hydrobiologia 397: 187-195.         [ Links ]
Villar C, Stripeikis J, Colautti D, D'Huicque L, Tudinio M, Bonetto C (2001). Metals contents in two fishes of different feeding behaviour in the lower Paraná River and Río de la Plata Estuary. Hydrobiologia 457: 225-233.         [ Links ]
Villar C, Tudinio M, Bonetto C, De Cabo L, Stripeikis J, D'Huicque L, Troccoli O (1998). Heavy metal concentrations in the Lower Paraná River and right margin of the Río de la Plata Estuary. Verh Internat Verein Limnol 26: 963-966.         [ Links ]
Villar CA, Gómez SE, Bentos CA (2000). Lethal concentration of Cu in the Neotropical fish Cnesterodon decemmaculatus (Pisces, Cyprinodontiformes) Bull Environ Contam Toxicol 65: 485-469.         [ Links ]
Welsh PG, Parrott JL, Dixon DG, Hodson PV, Spry DJ, Mierle G. (1996). Estimating acute copper toxicity to larval fathead minnow (Pimephales promelas) in soft water from measurements of dissolved organic carbon, calcium, and pH Can J Fish Aquat Sci 53: 1263-1271.         [ Links ]
Wetterhahn J (1979). Chromate metabolism in liver microsomes. Biol Trace Elem Res 1: 55-62.         [ Links ]
Wicklund Glynn A, Norregren L, Müssener Å (1994). Differences in uptake of inorganic mercury and cadmium in the gills of zebrafish, Brachydanio rerio. Aquat Toxicol 30: 13-26.         [ Links ]

Received on April 18, 2005.
Accepted on May 31, 2005.

Creative Commons License Todo el contenido de esta revista, excepto dónde está identificado, está bajo una Licencia Creative Commons