Analysis of Selected Herbicides And Related Metabolites in a Coastal Lagoon Under The Influence of Water Runoff: Sediment and Biota

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Coastal lagoons are subjected to strong anthropogenic pressures as a consequence of their location between Land and Sea. Among other contaminants, they are receiving high nutrient loadings and pesticides, mainly of agricultural origin, from heavily exploited watersheds.

Since the number of chemicals released into the water systems is high, there is a need to classify them and the associated risk, basically considering their environmental impacts and the deterioration of aquatic ecosystems from their adverse effects. For these reasons, the European Commission has produced a draft list of “priority hazardous substances”, which are of particular concern for waters and which “will be subject to cessation or phasing out of discharges, emissions and losses within an appropriate timetable that shall not exceed 20 years” (Directive 2000/60/EC, Decision 2455/2001/EC). Chemicals were selected on the basis of environmental hazard criteria, namely their toxicity, persistence and bioaccumulation. Monitoring activities and environmental risk assessment for substances in the priority list have been also reported (93/67/EEC and Commission Regulation No. 1488/94). Furthermore, the Decision 2455/2001/EC ranks in order of priority the substances for which environmental quality standards and emission control measures will be set at the European Union scale.

Among the plant protection products, atrazine and simazine have been included in the list, due to their persistence and toxicity features, their relatively high water solubility and leaching capacity and low potential to adsorb onto soils and sediment (Meakins et al.,1995). Their persistence or degradation depend upon several environmental factors, namely sunlight, temperature, pH, organic matter and microorganisms (Chung and Gu, 2003), whilst toxicity, bioaccumulation and biomagnification in the foodweb are mainly a function of physico-chemical properties of the pesticide molecule (Pérez- Ruzafa et al., 2000; Gluth et al., 1985).

Overall, pathways and fate of pesticides from agricultural sources through aquatic ecosystems are difficult to predict and depends upon meteorological conditions, the delivery being affected by leaching and runoff that may occur after strong rainfall events (Neumann et al., 2003).

Within the aquatic ecosystem, contaminants pathways are mainly driven by physical and physico-chemical processes. Among these, sorption onto suspended particulate matter and settling lead to a reduction of concentration and bioavailability in the water column (Voulvoulis et al., 2002). Thereafter, sediments can act as a sink with a
retention capacity which is related to factors and conditions mentioned above (Meakins et al.,1995). Among these, the sorption capacity results in a better contaminant degradation, since chemical reactivity and microbial processes attain their maximum at the solid-liquid interfaces. Within sediment, sorption and hydroxylation of contaminats are positively related to organic carbon and clay content. In this context, a low content of clay or organic matter results in less binding
sites, with a lower sorption capacity and with a subsequent increase in persistence. By contrast, contaminants can be trapped and retained by sediment particles with irreversible binding enhancing their persistence in deep anoxic sediment horizons. (Smalling and Aelion, 2006). Under these circumstances, sediments can act as a contaminant storage, preserving them for long times even after pesticides production and use are banned.

After their introduction in the environment, plant protection products are subject to several chemical and physical processes, which lead to degradation into metabolites, more or less toxic than the former molecules. Furthermore, some microrganisms are able to accumulate, detoxify or use organic contaminants as a substrate for their
metabolism. Chemical degradation consists of hydrolysis, redox, and photodecomposition reactions, which are, in general, well described by first order kinetics and controlled by several environmental factors like temperature, pH, solar radiation intensity, etc.

Pesticides with high vapour pressure and low water solubility such as organochlorines, are dispersed widespread, due to volatilization during spraying or after application, and dry/wet deposition (Oehme, 1991). Contaminants that show some resistance to degradation processes, if not adsorbed onto organic colloids in soil or volatilized, can be transported through leaching, drainage and diffusion towards ground waters and/or fresh and coastal waters. Here they may persist, depending on environmental conditions and on their chemical physical properties (e.g. solubility, Henry constant, half-life time, vapour pressure, octanol-water partition coefficient, organic carbon partition coefficient, etc.). These compounds and their metabolites, can reach aquatic ecosystems and can be bio-accumulated by biota through direct (respiration, epidermal absorption) and indirect (diet) uptake. Persistence of these contaminants in aquatic environment is extremely variable, in relation to degradation and fate-transport pathways, reported values ranged between few hours to several years.

Among contaminants from agricultural sources, herbicides are the most abundant. The different herbicide classes show behaviours which largely depend on environmental conditions and molecular properties. Agricultural practices and crop typologies are also relevant, since herbicides are delivered with frequencies that are related to crop cycles. As a consequence, the contamination timing in combination with altered environmental conditions (e.g., drought, anoxia, etc.) could enhance accumulation and persistence.

Among the most used and well known compounds, s-triazines undergo microbial degradation mainly in oxic conditions, whilst anaerobic decomposition is much less effective. For this reason, s-triazines are preserved in deep reducing sediment horizons, where long half-lives have been detected (Gu et al., 2003). Toxic effects of s-triazines on algae (e.g. Jianyi Ma et al., 2006; Podola and Melkonian, 2005; Dorigo and Leboulanger, 2001; Nitschke et al., 1999; Abdel- Hamid, 1996) acquatic plants (e.g. Cedergreen and Streibig, 2005) and aquatic microorganisms (De Lorenzo, 2001) have been reported. A complete review of most important studies on the effects of these contaminants can be found in Eisler (2000). A combined additive effect of s-triazines on algae has been also described by Faust et al. (2001) and Strachan et al. (2001). However, relatively few analytical field data are available for evaluating the toxic effects in marine fauna, especially for fishes and molluscs (Hall et al., 1994; Ward and Ballantine, 1985). Few data are available for atrazine in juvenile clam Mercenaria mercenaria, whose LOEC values are 1250 and 1000 μg L-1 in 96 h acute assay and 10 days chronic assay, respectively, with a NOEC of 500 μg L-1 (Lawton et al., in press). In this case sublethal endpoints were dry mass, shell size and condition index (dry mass/shell volume). Cheney et al. (1997) investigated the variation in metabolic activity in gill tissue of the mollusc Elliptio complanata in short time (20-50 min) exposure to atrazine in concentrations between 10-6 and 10-3 M. Finally, Losso et al. (2004) measured embryotoxicity (malformed larvae and pre-larval stages) of atrazine in early life stage of the bivalve Mitilus Galloprovincialis.

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