Chapter 3. Toxicity evaluation

Transcription

1 Chapter 3 Toxicity evaluation INTRODUCTION Toxicity can be defined as the inherent capacity of a toxicant to affect adversely any biological activity of an organism. The toxicity of an insecticide to an aquatic organism is usually expressed in terms of LC50 (Tripathi and Shukia, 1988). The best method for evaluating toxicity of a toxicant is by the determination of LC50 or LD50. This value represents the amount of a toxicant either in the form of lethal dose (LD) or concentration (LC50), which kills 50% of the population of the test animal with in a fixed period of time (Finney, 1971). If the test animals are terrestrial, the toxicant is administered generally either through the oral or intramuscular or inhalation methods and the toxicity is expressed in terms of LD50. Where as if the test animals are aquatic the toxicant will be generally mixed with the ambient medium and the toxicity is expressed in terms of LC50. The period of exposure is considerably important in evaluating the toxicity levels of toxicants in aquatic animals. Generally depending upon the nature of toxicant LC50 values are assessed at 24 or 48 or 96 hours or even more (Spehar, et. al, 1982; David and Philip, 2005). Aquatic toxicity test are performed in order to evaluate the response of aquatic organisms and to detect or measure the presence or effect of one or more substances, wastes or environmental factors, alone or in combination (Yazdandoost and Katdare, 1999). Evaluation of toxicity of chemical compounds is a complex phenomenon but yet a debating topic for scientists. 52

2 The bio-assessment of the toxicity of pesticides with reference to aquatic biota is playing a crucial role in establishing the toxicity evaluation with reference to aquatic fauna. The wide spread usage of synthetic insecticides ultimately pollute the aquatic environment there by affecting the aquatic fauna mainly fishes, which constitute the major economy of the country and valuable source of protein (Muniyan and Veeraragahavan, 1999). The complexities of the interaction of the pesticides with the biological systems of the fishes received greater attention of scientists. Acute toxicity is a major subject of research in all research institutes for evaluating the chemical toxicity test for assessing the potential hazards of chemical contamination to aquatic organisms (Muniyan and Veeraragahavan, 1999; Jain and Mishra, 1995). This concern has induced an increased awareness and closer scrutiny about the impact of these chemicals on fishes and fishery resources. Hence, investigation efforts of this nature would not only focus attention on conservation of fishes but also towards prevention of disorders caused to human beings by way of biomagnifications of these pollutants. The potent toxic exposure leads to interaction between the chemical agent and the target species in space and time. Apparently it must be characterized by dosage, type of exposure and characteristics of exposed populations. In general, the extent of biological effects of chemicals can be seen at molecular, cellular, tissue, organismic, family and population levels through behavioural, physiological and pathological means (Green, 1984; Patil and Dhande, 2001). 53

3 Recently, Schulz (2004) has reviewed existing knowledge of the exposure to and effects of insecticides in natural surface waters due to contamination from agricultural nonpoint-sources. The major routes of transport of insecticides, and other pesticides, from crop fields to adjacent streams are via surface run-off, drains, groundwater, wind drift and atmospheric deposition. Often the pesticide contamination occurs as transient pulses with elevated concentrations, especially when a field spraying is followed by significant amounts of precipitation or when pesticides are released as a result of illegal handling, e.g. direct over spraying or discharges from farmyards where pesticides are handled and spraying equipment is cleaned (Nørum, et. al, 2010) Although the effects of the pesticides in freshwater ecosystems remain to be fully elucidated, the concentrations observed, typically <10µgL 1, cannot be expected to be acutely lethal. Still, effects in streams at the population and ecosystem level following transient pulses of insecticides, especially pyrethroids, originating from the spraying of crop fields, have been reported (Schulz, 2004; Schulz and Liess, 1999a). The pyrethroids are the most important group of insecticides used and according to the annual pesticide statistics accounts for 89% of the insecticide treated crop area was sprayed with pyrethroids in 2008 (Anonymous, 2009). The toxicity of the pyrethroids is caused by their agonistic effect on voltage dependent sodium channels in the nervous system (Narahashi, 1996; Vijverberg and Vandenbercken, 1990). Synthetic pyrethroids have been introduced over the past two decades for 54

4 agricultural and domestic use as replacements for more toxic pesticides, such as chlorinated hydrocarbons, organophosphates and carbamates (Moore and Waring, 2001). The ecological implications of this large-scale shift in pesticide application have not yet been fully explored. Pyrethroids have reported to be extremely toxic to fish, some beneficial aquatic arthropods (for example lobster, shrimps, prawn, crabs, honey bees etc.,) in laboratory tests. Pyrethroids toxicity in mammals, birds, amphibians and both terrestrial and aquatic invertebrates has been comprehensively reviewed (Bradbury and coats, 1982). Cypermethrin is immobile and not expected to biomagnify through food chain. Due to its lipophilicity, pyrethroids have a high rate of gill absorption thereby rendering fish as most sensitive to the pesticides. Yet toxicity of pyrethroids to aquatic organisms varies widely depending upon its stereochemical structure (Milam, et. al, 2000) and its solubility in the diluent medium (Datta and Kaviraj, 2003). With reference to pyrethroids toxicity studies concerning osmotic stress, temperature, species, age, sex, ph and nutrition of test animals were also carried out to certain extent on aquatic organisms (Elliott, et. al, 1978; Jolly, et. al, 1978; Anderson, 1982; Khan, 1983; Ghosh and Chatterjee, 2002; Ravinder, 1988, Mc Kenney and Hamarker, 1984; Desai, et. al, 1986; Sanawane, 1999). Although volumes of available literature show the lethality to fishes, toxicity responses varies from species to species evidencing differences in their values. Much of the earlier work is inconsistent on fish toxicity and is 55

5 largely due to the differences in organism responses, substance used (commercial formulation, emulsified concentrate, technical and analytical grade) and experimental conditions. Therefore, it is pertinent to evaluate the susceptibility of the aquatic resources to cypermethrin. The main objective of this study was to determine the acute toxicity of technical grade cypermethrin to the freshwater teleost, Labeo rohita to assess potential risk of the cypermethrin. As these values are highly essential in evaluating the toxicity levels, hence the present investigation is commenced with the determination of 96 hours LC50 for cypemethrin to freshwater fish, L. rohita. RESULTS Studies to evaluate the toxicity of cypermethrin to Labeo rohita (3±0.5 g) exposed for 96 h showed nil mortality at 2.5 µg/l and 100% mortality was observed at 5.0 µg/l. The mortality rate increased with increase in the concentration of cypermethrin (Table.1). When percent mortality was plotted against log concentration of cypermethrin, a sigmoid curve was obtained (Fig. 1). The 96 h LC50 value was obtained from sigmoid curve was 4.0 µg/1. The percent mortality after transforming to probit mortality was plotted against log concentration of cypermethrin using probit method (Finney, 1971). In this a straight line was obtained and the LC50 value obtained from the graph was 4.0 µg/l also determined the 96 h LC50 value (Fig 2). The 96 h LC50 value was further verified by the method of Dragstcdt-Behren s equation (Carpenter, 1975) and the value calculated by this method was found to be 3.97 µg/l. Thus the average 96 h LC50 value determined by the above three methods was

6 µg/1 (Table 2). The upper and lower 95% confidence limits were found to be μg/l and μg/l, respectively (Table 3). DISCUSSION The acute test for a long time has been a major component in toxicity testing (Braunbeck, 2005). In which acute chemical toxicity is determined as a 96 h LC50 value. However the environmental significance of death of individuals after short term exposure to high concentration is questionable (Marigoudar, et. al, 2009). In the science of aquatic toxicology, fish play an important role in toxicity testing and hazard evaluation, as do the white rat and guinea pig in mammalian toxicology (Annon, 1972). The bio-assessments of toxicity of cypermethrin with reference to aquatic biota, especially fish is crucial in establishing the toxicity evaluation. Toxicity of pyrethroids to different fish species varies between ppm (Table 4). This data clearly indicates that cypermethrin is also toxic to Labeo rohita as to other fish species mentioned in the table, since cypermethrin is readily taken up by aquatic organisms. The 96 h LC50 value of cypermethrin to Labeo rohita was found to be 4.0 µg/1 (Table 3). Lethality in the present study is comparable to the few previously published studies that exist, for Rainbow trout 96 h LC50 ranges between µg/1, Salmo gairdaneri 8.0 µg/1, but that LC50 s exceeded this concentration. In laboratory tests cypermethrin was reported to be highly toxic for aquatic organisms with LC50 values ranging from 0.08 µg/l to 2 µg/1 for newly hatched shrimp. This can be attributed to the inability of the L. 57

7 rohita to withstand and metabolize the cypermethrin intoxication. The acute toxicity treatments showed strong negative effects on survival as pesticide concentration increased. This suggests dose-dependent survival and concentration graded lethality (Table. 1). The varying degree of mortality reported in this study is consistent with the earlier reports. It has been suggested that differences in an organism s biological adjustment and behaviour response to change in water chemistry (Prashanth and David, 2010). Synthetic pyrethroids are relatively safe for mammalian and avian species. Oral LD50 for rats and mice ranges from 100 to 200 mg/kg body weight (Casida, et. al, 1983). Acute oral toxicities of greater than 4000 have been reported for three pyrethroids in three species of birds (Bradbury and Coats, 1982; Smith and Stratton, 1986) indicating that avian species are also highly resistant to pyrethroid intoxication. Birds have a longer capacity to eliminate the pyrethroids than mammals, are themselves more efficient than fish, and they have difficulty in rapidly degrading pyrethroids. The high toxicity of pyrethroids in fish is partly due to poor ability to metabolize them. Owing to their high lipophilicity, cypermethrin was found to concentrate in the fat of fish (Bradbury, et. al, 1986). Various symptoms of poisoning can be observed from studies involving the determination of LC50. In the present study, the fish maintained in normal freshwater behaved in usual manner i.e., they were very active with their well co-ordinated movements. They were alert at slightest disturbance. 58

8 But at the sub lethal concentrations of cypermethrin they became irritable and hyper-excited. Jumping movements as well as restlessness were observed and finally the fish turned upside down. Mucus secretion and loss of equilibrium were also observed. They slowly became sluggish with short jerky movements, surfacing and gulping of air and erratic circular movements. Finally they settled down at the bottom with loss of equilibrium and rolling of the body, convulsions prior to death. The fish very often come to the surface in order to avoid toxic environment. Moreover, examination of the gill of dead fish revealed that the gill lamellae colour was changed from red to brown. Tilapia exposed to lethal and sub lethal concentration of endosulfun and lidane exhibited abnormal behaviour at lethal concentration; a sudden heavy stress on the fish showed erratic swimming, convulsion, spiraling, tremors, jerky movements and rapid opercular movements. The fish struggled hard for breathing, often moved to the surface to engulf atmospheric air and tried to escape the toxic aquatic medium. After a few hours, equilibrium was lost and the fishes spiraled and slowly moved upward in a vertical position. Finally they lost equilibrium completely and were flat at the bottom (Thorat, 2001). David (1995), Deva Prakasa Raju (2000), in Labeo rohita and T. mossambica respectively, also observed similar symptoms. Some workers have reported the toxicity evaluation of different pesticides in the same freshwater fish, Labeo rohita. The differences can be attributed either to potency of pesticide or difference in the test conditions. Further, may also be viewed as the quality/concentration of the test substance 59

9 (Emulsified, wettable powder, or technical grade) used. The unusual behaviour of the fish Labeo rohita in stress condition may be due to obstructed functions of neurotransmitters. The gill opercular movements increased initially to support enhanced physiological activities in stressful habitat and later decreased activities may be due to mucus accumulation on the gill. The toxic stress of pesticides has direct bearing on tissue chemical compounds (Tilak and Yacobu, 2002). This was also reported by Koundinya and Ramamurthy, 1980, Rajamanickam and Karpagaganapathy, 1988, Prabhakar, et. al, 1993, David, 1995 and Muniyan and Veeraragahavan, 1999, Chaudhary, et. al, The excessive secretion of mucus over the gills may inhibit the diffusion of oxygen during the process of gaseous exchange. It suggests that the cypermethrin is not safe to non-target organisms like fishes. It is concluded that cypermethrin is innately toxic to fish, Labeo rohita. Moreover, it is being considered important to select the species which can represent the particular locality. Since, toxicity responses or expressions differ from species to species and native conditions. The above-analogy also warrants for an indispensable need to evaluate more toxicity data for wide range of animal groups of the eco-web in order to understand the broad spectrum of cypermethrin in comparison to other pesticides available. This also provides a platform to establish tolerable limits and safe levels of toxic agents for the biota of aquatic environment and to save the residue imbalance in aquatic bio-ecological cycles, which help in involving bio-detector monitoring. Hence this type of study can be useful to compare the sensitivity of the various species of aquatic animals and potency of chemicals using LC50 60

10 values and to derive safe environmental concentration, by which there is no lethality and stress to the animals. Table 1: Mortality of Labeo rohita in different concentration of cypermethrin at 96 hours of exposure period Sl. No Concentration of Cypermethrin (µg/l) Log concentration No of fish exposed No of fish alive No of fish dead Percent mortality Probit mortality

11 Table 2: LC 50 value of cypermethrin for Labeo rohita after 96 hours of exposure Sl. No. Name of the Method LC50 value (in µgram) 1 Percent mortality (Sigmoid curve) 4.0 µg/l 2 Probit mortality (Linear curve) 4.0 µg/l 3 Dragsted and Behren s method 3.97 µg/l Mean 3.99 µg/l Standard Deviation

12 Table 3: 96 h LC50, slope and 95% confidence limits of cypermethrin in the freshwater carp, Labeo rohita Pesticide 96 h LC50 value (μg/l) Slope 95% Confidence limits Upper limit Lower limit Cypermethrin 4.0 ±

13 Table 4: Data showing the acute toxicity of different pyrethroids on fish species Test Species Pyrethroid Exposure period (h) Concentration of Pyrethroid (µg/l) Reference O. kisutch Allethrin Mauck, et.al., 1976 Salmo gairdeneri Allethrin Mauck, et. al., 1976 Oryzias latipes Permethrin Miyamoto, 1976 O. kisutch Permethrin Mauck, et. al., 1976 L. macrochirous Permethrin Mauck, et. al., 1976 Salmo gairdneri Permethrin Mulla, et. al., 1978 Gambusia affinis Permethrin Mulla, et. al., 1978 Salmo gairdneri Permethrin Holcombe, et. al.,1982 C. veriegatus Permethrin Schimmel, et. al., 1983 Oryzias latipes Tetramethrein Miyamoto, 1976 L. macrochirus Tetramethrein Worthing, 1983 Salmo salar Deltamethrin Zitko, et. al., 1974 Gambusia offinis Deltamethrin Bocquet, 1985 C. mrigala Deltamethrin Naik, 2009 L. variegates Flucythrirate Schimmel, et. al., 1983 T. mossambica Fenvalerate Radhaiah,1988 C. mrigala Fenvalerate 21 days Sheela, et. al., 1992 Channa striatus Fenvalerate 21 days Sheela, et. al., 1992 Labeo rohita Fenvalerate David, 1995 C. mrigala Fenvalerate Anita, et. al., 1999 T. mossambica Fenvalerate Deva Prakasa Raju, 2000 C. idellus Fenvalerate Tilak, et. al., 2001 C. idellus Fenvalerate Tilak, et. al., 2002 Labeo rohita Fenvalerate µg/l David, et. al.,

14 Table 5: Data showing the acute toxicity of cypermethrin to different species of fishes Test Species Exposure period (h) Concentrat ion (µg/l) Reference Rainbow trout Coats and Donnel, 1979 Salmo salar Mc Leese, et. al, 1980 Rainbow trout Chapman, et. al, 1981 Cyprinus carpio Stephenson, 1982 Salmo gairdeneri Stephenson, 1982 Salmo gairdeneri Edwards, et. al, 1986 Cyprinus carpio Malla Reddy, 1987 Cyprinus carpio Malla Reddy, 1989 L. thermalis Jebakumar, et. al, 1990 Labeo rohita Sridevi, 1991 Tilapia mossambica 96 9 Reddy and Yellamma, 1991 Labeo rohita 48 6 Malla Reddy, et. al, 1989 Cyprinus carpio 21 day 3 Piska et al, 1992 Cyprinus carpio Malla Reddy, et. al,1995 Cyprinus carpio Sivakumari, et. al, 1997 Heteropneustes fossilis Bhargava and Rawat, 1999 C.mrigala µg/l Prashanth,

15 Percent mortality Fig 1: Toxicity evaluation of cypermethrin to freshwater fish, Labeo rohita. The graph showing sigmoid curve between percent mortality of fish against log concentration Log conc. 66

16 Probit kill Fig. 2: Toxicity evaluation of cypermethrin to freshwater fish, Labeo rohita. The graph showing linear curve between probit mortality of fish against log concentration Log conc 67