In the present study, specimens of Clarias gariepinus having mean weight and standard length of 197.38 ± 2.34 g and 27.36 ± 0.23 cm respectively were exposed to an organophosphate pesticide, fenthion. The median lethal concentration value (LC50) of fenthion and the effects of sub-lethal concentrations of the pesticide on the oxidative stress biomarkers, haematological parameters and histopathological changes in the blood and tissues of Clarias gariepinus were determined. The 24, 48, 72 and 96 hour LC50 values (with 95% confidence limits) estimated by probit analysis were found to be 61.987, 50.318, 45.056 and 39.968 mg/L, respectively. Significant differences (p < 0.05) at different durations of LC50 were observed. A concentration-dependent increase and time-dependent decrease was observed in mortality rate. Behavioural responses such as hyperactivity, erratic swimming, skin discoloration, jerky movement, followed by exhaustion and death were observed during the acute toxicity test. Fish were exposed to 2.0, 4.0 and 8.0 mg/L corresponding to 1/20th, 1/10th and 1/5th respectively, of 96 hour LC50 of fenthion for 21 days and 7 days recovery to evaluate the haematological parameters, enzyme activities and histopathological changes. There was no significant difference between the condition factors of the treated fish and the control (p > 0.05) but values of hepatosomatic index of the treated groups were reduced compared to the control though the reduction was not significant (p > 0.05) except on day 21 of the exposure. The liver and gill tissues were sampled on day 1, 7, 21 and during the 7 days recovery for analysis of the different parameters. Induction of oxidative stress in the liver and gill tissues of Clarias gariepinus were evidenced by increased lipid peroxidation (LPO) levels (p < 0.05), such that increased LPO was higher in fish exposed to the highest concentration of fenthion. The values of glutathione (GSH) activities were not significant (p > 0.05) except on day 1 of the exposure where there was significant increase (p < 0.05) in GSH activities in liver of fish exposed to the different concentrations of fenthion compared to the gills. Slight decrease in the activity of glutathione reductase (GR) was observed in the liver and gill tissues of the treated fish compared to the control but the reduction was not significant (p > 0.05). Increase in GR activities was observed in post-exposure (7 days recovery). There was a general reduction in glutathione peroxidase (GPx) activities but not significant (p > 0.05) except on days 1 and 7 where liver tissues exposed to the highest concentration of fenthion differed significantly compared to the gill tissues (p < 0.05). Increase in catalase (CAT) activities was both time and concentration dependent in all the tissues. CAT activity increased significantly in the liver of fish exposed to the highest concentration compared to gill tissues (p < 0.05) except on day 7. Superoxide dismutase (SOD) activities in the liver decreased from day 1 of the exposure to day 7 (though not significant). SOD activity slightly increased from day 14 of the exposure to post exposure (though not significant) except on day 21 where the liver tissues significantly increased compared to the gill tissues (p < 0.05). The exposure of fish to fenthion revealed that the pesticide had adverse effects on various blood parameters. Red blood cell (RBC) counts of the treated fish were lower compared to the control (p < 0.05). Packed cell volume (PCV) of the treated fish were lower than that of the control but the difference was not significant (p > 0.05) except in fish exposed to 8.0 mg/L on day 1 of the exposure (p < 0.05). Haemoglobin concentration (Hb) in fish decreased after the exposure though not significant (p > 0.05) but fish exposed to the highest concentration had significant increase in Hb at days 1 and 21 of the exposure. Significant increase in white blood cell (WBC) counts was also observed (p < 0.05). No significant difference (p > 0.05) in erythrocyte indices, mean corpuscular volume (MCV), mean corpuscular haemoglobin concentration (MCHC) and mean corpuscular haemoglobin (MCH) was observed in all the concentrations during the experiment. Changes in white blood cell differentials (neutrophils, lymphocytes, monocytes, eosinophils and basophils) revealed that no significant difference (p > 0.05) was observed in the parameters except in monocytes. The Histopathological effects of fenthion on the liver and gill tissues were also studied. The liver and gill tissues of the control exhibited quite a normal architecture. The main histopathological changes caused by fenthion in the liver of fish exposed to different concentrations were disarray of hepatic chords, vacuolation, sinusoid enlargement and necrosis. Liver suffered from toxicity as hypertrophy of hepatocytes and necrosis. These hepatic alterations were more evident in fish exposed to the highest concentration of fenthion. Alterations in gill structure exposed to the highest concentration were oedema, lifting of lamellar epithelia, destruction of gill architecture and lamellar fusion. Gill disorder and fusion of the secondary lamellar were pronounced in all treatments. The present study revealed that sub-lethal concentrations of fenthion resulted in the alterations of the studied parameters. Histopathogical changes observed during the period of the experiment as well as the recovery shows that fenthion-induced oxidative stress is irreversible.
TABLE OF CONTENTS
Title page – – – – – – – – – – – i
Certification – – – – – – – – – – – ii
Dedication – – – – – – – – – – – iii
Acknowledgements – – – – – – – – – – iv
Abstract – – – – – – – – – – – v
List of abbreviations – – – – – – – – – – vii
Table of contents – – – – – – – – – – ix
List of tables – – – – – – – – – – – xiv
List of figures – – – – – – – – – – – xvi
List of plates – – – – – – – – – – – xvii
CHAPTER ONE: INTRODUCTION AND LITERATURE REVIEW
1.1 Introduction – – – – – – – – – – 1
1.1.2 Statement of the problem – – – – – – – – 5
1.1.3 Justification of the study – – – – – – – – 6
1.1.4 Objectives of the study – – – – – – – – 7
1.2 Literature Review – – – – – – – – – 9
1.2.1 Pesticides – – – – – – – – – – 9
1.2.2 Organophosphate insecticides – – – – – – – – 10
1.2.3 History and composition of organophosphate compounds – – – 12
1.2.4 Pesticide metabolism – – – – – – – – – 14
1.2.5 Mechanism of action of organophosphate pesticides – – – – – 17
1.2.6 Oxidative stress induced by pesticides – – – – – – 19
1.2.7 Organophosphate pesticides, oxidative stress and antioxidant status – – 20
1.2.8 Fenthion – – – – – – – – – – 21
1.2.9 Description of clariidae – – – – – – – – 23
CHAPTER TWO: MATERIALS AND METHODS
2.1 Experimental Fish Management and Chemicals – – – – – 26
2.2 Determination of Median Lethal Concentration (Acute Toxicity),
NOEC and LOEC – – – – – – – – – 26
2.2.1 Determination of median lethal concentration – – – – – 26
2.2.2 Determination of no observed effect concentration (NOEC) – – – – 27
2.2.3 Determination of lowest observed effect concentration (LOEC) – – – 27
2.3 Determination of Safe Levels – – – – – – – – 27
2.4 Observation of Behavioural Responses – – – – – – 28
2.5 Experimental Design for Chronic Exposure – – – – – – 28
2.6 Lipid Peroxidation – – – – – – – – – 29
2.7 Changes in Activities of Antioxidant Enzymes – – – – – 30
2.7.1 Catalase activity – – – – – – – – – 30
2.7.2 Superoxide dismutase – – – – – – – – – 30
2.7.3 Reduced glutathione – – – – – – – – – 31
2.7.4 Glutathione reductase – – – – – – – – 31
2.7.5 Glutathione peroxidase – – – – – – – – 32
2.8 Haematological Parameters – – – – – – – – 32
2.8.1 Determination of packed cell volume – – – – – – – 32
2.8.2 Determination of haemoglobin – – – – – – – 32
2.8.3 Determination of red blood cell count – – – – – – 33
2.8.4 Determination of white blood cell count – – – – – – 33
2.9 Condition Factor – – – – – – – – – – 34
2.10 Hepatosomatic index – – – – – – – – – 35
2.11 Physico-chemical Parameters of the Test Water – – – – – 35
2.11.1 Measurement of water temperature – – – – – – – 35
2.11.2 Determination of dissolved oxygen – – – – – – – 35
2.11.3 Determination of pH – – – – – – – – – 36
2.11.4 Determination of turbidity – – – – – – – – 37
2.11.5 Determination of ammonia – – – – – – – – 37
2.11.6 Determination of nitrate – – – – – – – – 38
2.11.7 Determination of phosphate – – – – – – – 39
2.12 Histopathological Examination – – – – – – – 39
2.13 Statistical Analysis – – – – – – – – – 40
CHAPTER THREE: RESULTS
3.1 Median lethal Concentration (LC50), NOEC and LOEC – – – – 41
3.1.1 Median lethal concentration (LC50) – – – – – – – 41
3.1.2 Statistical endpoints of acute toxicity test – – – – – – 43
3.2 Estimation of Safe Levels – – – – – – – – 46
3.3 Behavioural Responses of Clarias gariepinus Exposed to Fenthion – – – 48
3.4 Effects of Fenthion on Lipid Peroxidation (MDA) in the Liver and Gill
Tissues of Clarias gariepinus– – – – – – – 50
3.5 Effects of Different Concentrations of Fenthion on Antioxidant
Enzymes in Clarias gariepinus – – – – – – – 52
3.5.1 Effects of fenthion on catalase (CAT) activity in the liver and gill
tissues of Clarias gariepinus – – – – – – – – 52
3.5.2 Effects of fenthion on superoxide dismutase (SOD) activity in the liver and
gill tissues of Clarias gariepinus – – – – – – – 54
3.5.3 Effects of fenthion on reduced glutathione (GSH) activity in Clarias gariepinus – 56
3.5.4 Effects of fenthion on glutathione reductase (GR) activity in the liver and
gill tissues of Clarias gariepinus – – – – – – – 58
3.5.5 Effects of fenthion on glutathione peroxidase (GPx) activity in the liver and
gill tissues of Clarias gariepinus – – – – – – – 60
3.6 Haematological Parameters – – – – – – – – 62
3.6.1 Effects of exposure to sub-lethal concentrations of fenthion on
Packed cell volume (PCV) in Clarias gariepinus – – – – – 62
3.6.2 Haemoglobin (Hb) changes in Clarias gariepinus exposed to
sub-lethal concentrations of fenthion – – – – – – – 64
3.6.3 Effects of exposure to sub-lethal fenthion concentrations on
red blood cells in Clarias gariepinus – – – – – – – 66
3.6.4 Effects of exposure to sub-lethal concentrations of fenthion on
white blood cells in Clarias gariepinus – – – – – – 68
3.6.5 Effect of exposure to sub-lethal fenthion concentrations on mean
corpuscular haemoglobin concentration (MCHC) in Clarias gariepinus – – 70
3.6.6 Effect of exposure to sub-lethal fenthion concentrations on mean
corpuscular haemoglobin (MCH) in Clarias gariepinus – – – – 72
3.6.7 Effect of exposure to sub-lethal fenthion concentrations on mean
corpuscular volume (MCV) in Clarias gariepinus – – – – 74
3.6.8 Effect of exposure to sub-lethal fenthion concentrations on
neutrophils in Clarias gariepinus – – – – – – – 76
3.6.9 Effects of exposure to sub-lethal fenthion concentrations on
lymphocytes in Clarias gariepinus – – – – – – 78
3.6.10 Effects of exposure to sub-lethal fenthion concentrations on
monocytes in Clarias gariepinus – – – – – – – 80
3.6.11 Effect of exposure to sub-lethal fenthion concentrations on
eosinophils in Clarias gariepinus – – – – – – – 82
3.6.12 Effect of exposure to sub-lethal fenthion concentrations on
basophils in Clarias gariepinus – – – – – – – 84
3.7 Effect of exposure to sub-lethal fenthion concentrations on
condition factor in Clarias gariepinus – – – – – – 86
3.8 Effect of exposure to sub-lethal fenthion concentrations on
hepatosomatic index in Clarias gariepinus – – – – – – 88
3.9 Physico-chemical Parameters of the Test Water – – – – – 90
3.9.1 Effect of exposure to sub-lethal fenthion concentrations on
temperature of the test water – – – – – – – – 90
3.9.2 Effect of exposure to sub-lethal fenthion concentrations on
dissolved oxygen of the test water – – – – – – – 92
3.9.3 Effect of exposure to sub-lethal fenthion concentrations on pH of
the test water – – – – – – – – – 94
3.9.4 Effect of exposure to sub-lethal fenthion concentrations on turbidity
of the test water – – – – – – – – – 96
3.9.5 Effect of exposure to sub-lethal fenthion concentrations on ammonia
concentration of the test water – – – – – – – 98
3.9.6 Effect of exposure to sub-lethal fenthion concentrations on nitrate
concentration of the test water – – – – – – – 100
3.9.7 Effect of exposure to sub-lethal fenthion concentrations in
phosphate of the test water – – – – – – – – 102
3.10 Histopathological Results – – – – – – – – 104
3.10.1 Histopathological observations of gills in Clarias gariepinus
exposed to fenthion – – – – – – – – – 104
3.10.2 Histopathological observations of liver in Clarias gariepinus
exposed to fenthion – – – – – – – – – 115
CHAPTER FOUR: DISCUSSION, CONCLUSION AND RECOMMENDATIONS
4.1 Discussion – – – – – – – – – – 125
4.2 Conclusion – – – – – – – – – – 141
4.3 Recommendations – – – – – – – – – 143
References – – – – – – – – – – – 144
INTRODUCTION AND LITERATURE REVIEW
Living systems encounter a variety of stresses during their continuous interaction with the environment. Environmentally-induced stresses frequently activate the endogenous production of reactive oxygen species (ROS), most of which are generated as side products of tissue respiration. Hence, constant exposure to stressors may enhance ROS-mediated oxidative damage. Increased number of agricultural and industrial wastes enter aquatic environment and cause changes in aquatic organisms (Stoliar and Lushchak, 2012). Some of them directly enhance ROS formation whereas others act indirectly, for example, by binding with cellular thiols and reducing antioxidant potential.
Fresh water fish diversity is threatened by a number of environmental stressors including contaminants and nutrient loading, overharvesting, habitat degradation and climate change (Dudgeon et al., 2006; Jelks et al., 2008). The discharge of pesticides into water bodies is of accelerating concern and is likely to continue and possibly increase in the near future (Lucero et al., 2000). The use of pesticides has been recognized as part of agricultural practices globally (Nwani et al., 2011). For centuries, pesticides have been used to enhance the production of food by controlling disease vectors (Amin and Hashem, 2012). Although the use of pesticides is of great importance, there are problems associated with it (Quinn, 2007). The aquatic organisms are vulnerable to contamination as run-offs from farms and industries end up in water bodies (Botelho et al., 2009) and accumulate as huge amount of residues in the environment, thereby causing a substantial hazard in the environment due to its uptake and accumulation in the food chain and drinking water. The washing of packaging materials and application equipments often carried out on the banks of water bodies help in scaling up their contamination potential (Trovo et al., 2005). Furthermore, residues from pesticides and associated human activities such as deforestation, release of domestic, hospital and industrial effluents may contribute to a large buildup and discharge into the aquatic environment (Nwani et al., 2014). In the water, the molecules of pesticides may bind to the materials in suspension, accumulate in the sediment or pose a considerable risk to non-target organisms (Miles and Pfeuffer, 1997). They could also be absorbed by these animals through inhalation, orally or skin absorption with attendant physiological responses that may have effect on behaviour, morphological, enzymological, biochemical and antioxidant responses (Jordan et al., 2013).
Fish are particularly threatened by water pollution even though they have antioxidant defense system. Upon exposure to pollutants, free radicals such as superoxide (O2–), hydroxyl radicals (OH ⃰ ) and hydrogen peroxide (H2O2) are generated (Nwani et al., 2014). Under normal circumstances, there is a balance between the amount of free radicals generated in the body and the antioxidant defence systems that are able to cater for these free radicals (Kiew and Don, 2012). Problems would only arise when the amount of free radicals is beyond the normal physiological level induced by the environmental condition or produced within the body (Nose, 2002). The free radicals at excess reacts with biological macromolecules to increase the level of lipid peroxidation, protein denaturation and alterations in the activities of antioxidant enzymes such as catalase (CAT), superoxide dismutase (SOD), Glutathione reductase (GR) and Glutathione Peroxidase (GPx). These antioxidant enzymes are involved to counteract the toxicity of ROS (Orbea et al., 2002) thereby preventing the cells and tissues from oxidative damage. Excessive ROS offset the antioxidant capacity and build up oxidative stress. Pesticide-induced oxidative stress has been a focus of toxicological research in the last decade as a possible mechanism of toxicity (Akhgari et al., 2003; Ciccheti and Argentin, 2003; Abdollahi et al., 2004). Several circumstances promote the antioxidant defence responses in fish (Trendazor et al., 2006). Factors intrinsic to the fish itself as well as environmental factors such as type of diet, daily or seasonal changes in temperature, dissolved oxygen, toxins present in the water, pathogens or parasites can fortify or weaken antioxidant defences (Felton, 1995; Martinez-Alvarez, et al., 2005).
Haematological alteration is a good method for rapid evaluation of the chronic toxicities of compounds (Jaya and Ajay, 2014). A thin epithelial membrane separates fish blood from the water and any unfavourable change in the water body is reflected in the blood (Kori-Siakpere and Ubogu, 2008). Therefore, observations of different haematological parameters would provide a better understanding in diagnosing the effect of environmental stress on an animal and in fact, would give an insight into changes induced in the circulating fluid thereby offering a valuable knowledge on the correct monitoring of fish health status (Lazar and Lazar, 2012).
Histopathological studies could help to establish causal relations between contaminant exposure and various biological responses (Boran et al., 2012) and have proved to be a sensitive tool used in detecting direct effects of chemical compounds within target organs of fish in laboratory experiments (Altino and Capkin, 2007; Capkin et al., 2009). An organ can function normal only when its structure is normal but any structural damage to it is likely to affect the function of that organ (Muralidharan, 2014b). A study on histopathology would provide a very important and useful data concerning changes in cellular or sub cellular structure of an organ much earlier than external notification. One of the advantages of using histopathological biomarkers in environmental studies is that it allows the examination of target organs (Gernhofer et al., 2001). Also, the alterations found in these organs are easier to identify than functional ones (Fanta et al., 2003). These alterations serve as warning signs of damage to the well-being of an organism. Studies have been conducted on histopathological changes in the gills, liver and kidney of fish exposed to pesticides which have been reported to cause pathological alterations in the exposed fish (Auta, 2001). Histopathological changes have been used as important biomarkers in environmental monitoring that allows examining specific target organs (Devi and Mishra, 2013) such as liver and the gills.
Fenthion is one of the most widely used organophosphate insecticides and avicide in agriculture and public health for controlling many sucking and biting pests (Cong et al., 2009; Sevgiler and Uner, 2010). While it is effective as an insecticide, it is moderately toxic to mammals and highly toxic to birds through dermal contact and inhalation (Vlastos and Ganidi, 2004). Fenthion is characterized as a cholinesterase inhibitor as well as a lipophylic compound that is bioaccumulated in the fat tissues of animals (Vlastos and Ganidi, 2004). According to Environmental Protection Agency (2003) and Pest Management Regulatory Agency (2004), all Fenthion formulations have been banned in the United States and Canada. However, it is still produced in some countries such as China and India and the application of these insecticides is ongoing in Nigeria.
Biomarkers are key molecular or cellular events that link a specific environmental exposure to a health outcome (National Institute of Environmental Health, 2014). In other words, they are measurable indicators of some biological state or condition. The use of fish as biomarkers for assessing the effect of pollution is of increasing importance and permits early detection of aquatic environmental problems (Lorez-Barea, 1990; Van-Der Oost et al., 2003). Studies on pesticide-induced effects on various antioxidant enzymes in fish and other aquatic organisms can provide useful information about the ecological consequences of pesticide use (Kavitha Venkateswara-Rao, 2008). Gills are important in respiration, osmoregulation, acid-base balance and excretion of nitrogenous wastes in fish and they serve as the first area of contact of the animal with the external environment (Deb and Das, 2013). The liver is the main organ of detoxification due to its lipophilicity (Ali et al., 2009). Therefore, these tissues can be used in determining biomarkers of oxidative stress such as changes in antioxidant enzymes activity or the degree of accumulation of damaged molecules. This can offer an early warning sign for exposure to redox-active xenobiotics. Enzyme analyses are becoming increasingly important in the determination of toxic effects of chemical pollutants in environmental toxicology (Muralidharan, 2014a) since alterations in enzyme activity with respect to environmental change are used as stress indicators.
1.1.2 Statement of the problem
Aquatic environments are a home to a vast diversity of organisms ranging from prokaryotes to higher vertebrates. They are also of great importance to humans, providing essentials such as water and food as well as transportation, economic opportunities, recreation, etc. Unfortunately, these environments also act as sinks for a great variety of anthropogenic pollutants, many of which are toxic. Environmental pollution resulting from industrial effluents and agricultural activities has become a global issue due to the extent of damage caused to the aquatic organisms as the orderly balance in their physiological process is constantly under attack by the environmental adversities and the disruption in the natural food chain (Amin and Hashem, 2012; Nwani et al., 2013a; Muralidharan, 2014a). Most of the studies on oxidative stress in fish focused on toxicological aspects, such as effects of xenobiotics on antioxidant-enzyme activities, induction of biotransformation processes as well as on the intensity of lipid peroxidation and other biomarkers of oxidative damage (Winston and Giulio, 1991). Despite these parameters having been used as biomarkers for contaminants, a review of related literature revealed no clear trend since fish response depends on several variables such as the species, tissues, the antioxidant parameters itself, time of exposure and contaminant concentration, besides the other physiological and environmental changes mentioned above.
Studies have demonstrated that fenthion induce DNA damage and should be considered potentially hazardous to humans (Wu et al., 2011). Recent research by Kanter and Celik (2012) revealed that exposure of frogs to fenthion induce an increase in melondialdehyde (MDA), and antioxidant defence systems (ADS). Israel and Sam (2012) reported biochemical changes in some tissues of Cirrhina mrigala (Hamilton) (Cyprinidae: Cypriniformes) exposed to fenthion. Altuntas and Delibas (2002) reported that fenthion caused liver damage in Wister albino rats and may be one of the molecular mechanisms involved in fenthion-induced toxicity. Extensive works on the effect of fenthion on liver function of Cyprinus carpio have been performed (Gopi, 1993; Kitamure et al., 2003). The pigments of Cyprinus carpio exposed to fenthion have been studied (Muralidharan and Pillai, 2012). However, basic research that provides evidence for the oxidative stress biomarkers, haematological parameters and histopathological changes in the African catfish, Clarias gariepinus exposed to fenthion is lacking.
1.1.3 Justification of the study
The need to improve agriculture and public health standards has necessitated the widespread use of pesticides but these pesticides are drained into the natural habitat of fishes. Many of the pesticides are not biodegradable and due to accumulation, can enter into food chain and ultimately affect human and animal health as some of them are toxic. It is believed that fish possess the same biochemical pathways to deal with the toxic effects of endogenous and exogenous agents as do mammalian species (Al-Akel et al., 2010; Alkahem et al., 2011). Therefore, it is important to study the toxic effects of pesticides on fish since they constitute an important link in food chain and their contamination by pesticides affect the aquatic system. The study attempts to investigate the oxidative stress biomarkers, haematological parameters and histopathological changes in Clarias gariepinus exposed to fenthion as studies on oxidative stress in fish have opened a number of research lines on fish physiology in recent years. The study will provide information concerning the response of antioxidant defenses, haematological alterations and histopathological changes under different circumstances as well as regulatory mechanisms of these responses which will no doubt benefit aspects related to fish farming and production. The findings will also be beneficial as it will serve as a baseline for data assessment of health status of the catfish, Clarias gariepinus as well as provide governing bodies with information useful in management of aquatic ecosystems, information on which to base regulations concerning usage and handling of chemical compounds and essentially help to protect aquatic ecosystems from the negative effects of anthropogenic activities.
1.1.4 Objectives of the study
The objectives of this study are to;
- determine the median lethal concentration (LC50); no observed effect concentration (NOEC) and lowest observed effect concentration (LOEC) of fenthion in Clarias gariepinus.
- determine the safe levels of fenthion in gariepinus.
- observe the behavioural responses of gariepinus associated with fenthion exposure.
- investigate the effects of fenthion on lipid peroxidation in gariepinus.
- investigate the effects of fenthion on the activities of some selected antioxidant enzymes as biomarkers of oxidative stress in different tissues of gariepinus.
- investigate the changes in haematological parameters in gariepinus exposed to fenthion.
- evaluate the changes in condition factor and hepatosomatic index of Clarias gariepinus exposed to fenthion.
- assess the changes in physico-chemical parameters of the test water.
- investigate the histopathological changes in the gill and liver tissues of Clarias gariepinus exposed to fenthion.
1.2 LITERATURE REVIEW
Pesticides are organic compounds manufactured and used for the control of pests (Ortiz-Hermandez and Sanchez-Salines, 2010). Pesticides can be defined as any chemical substance or mixture of substances intended for preventing, destroying, repelling or mitigating effects of any pest of plants and animals (Agarry et al., 2013). The main groups of pesticides are fungicides, herbicides, weedicides, rodenticides and insecticides (Sharma et al., 2014). Among these classes of pesticides, insecticides are widely used in agriculture to control insect pests. Pesticides are categorized based on their chemical structures (Sternerson, 2004) and the various categories are often used against a number of pests (Quinn, 2007) or weeds in modern agriculture technology for high yield and management of public health (Yadav et al., 2010). Some of the categories include; organochlorine, organophosphate, carbamate, synthetic pyrethroids, among others (Natala and Ochoje, 2009; Naqvi et al., 2011; Nsikak and Aruwojoye, 2011). Pesticides enter water bodies directly or indirectly resulting in contamination of various aquatic ecosystems.
The effects of insecticide pollution on non-target organisms in the environment can be studied by detecting changes in organisms at the physiological, biochemical or molecular levels, providing “early warning” signs in monitoring environment quality (Sharma et al., 2014). These sensitive early warning biomarkers measure interactions between environmental xenobiotics and biological effects. Inhibition and induction of these biomarkers is a reliable approach of measuring potential impacts of pollutants on environmental organisms (Rendo `n-von Osten et al., 2005).
The number of pesticide in use has been on the increase since World War II (Alavi et al., 2008). In Nigeria, the use of pesticide has been on the increase ever since its introduction in the early fifties for cocoa production (Agarry et al., 2013). It has been estimated that about 125,000 to 130,000 metric tons of pesticides are applied every year in Nigeria (Asogwa and Donga, 2009). Many of these compounds will linger in the environment for many years to come due to their environmental persistence. Humans as well as animals are inevitably exposed to pesticides, through occupational use or environmental contamination. Pesticides act selectively against certain organisms without adversely affecting others. Absolute selectivity, however, is difficult to achieve (Bolognesi, 2003).
1.2.2 Organophosphate (OP) insecticides
OP compounds have largely been used as pesticides in many parts of the world. These chemicals are readily available because of insufficient regulations to control their sale (Jeyarathnam, 1990).
OP are phosphorus-containing insecticides whose insecticidal qualities were first observed in Germany during World War II in the study of the extremely toxic OP nerve gases (sarin, soman, and tabun) (Cabello et al., 2001). The use of OP increased not only in agricultural environments (Fensake et al., 2002; Koch et al., 2002) but also significantly in residential and urban settings (Lu et al., 2001; Berkowitz et al., 2003). OP insecticides have been considered as alternatives to organochlorine insecticides due to their broad-spectrum pesticidal properties and relatively shorter persistence after applications (Sharma et al., 2005). OP agents in addition to their intended effects like control of insects or other pests are sometimes found even to affect non-target organisms including humans (Chaudhuri et al., 1999; Burns et al., 2013).
The effects of various OP insecticides in humans and animals include cholinergic and non-cholinergic biological disturbances (Gordon and Mack, 2003; Jain et al., 2009). The widespread use of organophosphate insecticides (OPIs) has long been shown to exert deleterious effects on living organisms (Gultekin et al., 2001) as they are powerful inhibitors of acetylcholinesterase (AChE). AChE is an enzyme that breaks the neurotransmitter acetylcholine (ACh), which activates choligenic neurons, the nerve cell that control the signals in the peripheral nervous system, brain and spinal cord (Bakry et al., 2006). If ACh is not inactivated immediately after it has done its job, the neuron becomes over stimulated leading to increased secretions, sensory and behavioural disturbances, uncoordination, depressed motor function, respiratory depression, convulsion and death (Reigart and Roberts, 1999). Recovery depends on the manufacture of more AChE. AChE is also involved in various clinical effects such as neck muscle weakness and diarrhoea which can occur due to organophosphate poisoning in humans (Serdar and Gibson, 1985; Grimsley et al., 1997). Measurement of the activity of this enzyme in aquatic animals does not only offer a means of detecting serious pollution by AChE agents but has the potential for indicating the extent of poisoning the animals. Through inhibition of AChE, OP poisoning is characterized by the clinical picture of acute cholinergic crisis. Other manifestations are intermediate neurotoxic syndrome and delayed polyneuropathy (Aardema et al., 2008). At high doses, OPs are irreversible inhibitors of AChE, producing accumulation of acetylcholine in cholinergic nervous system. However, little is known regarding possible toxic effects of exposure to low doses of OP compounds (Ceh and Majdic, 2010). Thus, in addition to AChE inhibition being the principal mode of action of OP pesticides, increased LPO has also been implicated in mediating OP-induced toxicity in animals (Kaur and Dhanju, 2004).
1.2.3 History and composition of organophosphate (OP) compounds
OP compounds were first developed by Schrader shortly before and during the Second World War. They were first used as an agricultural insecticide and later as potential chemical warfare agents (Taylor, 1996). In the late 1990 and 2000, with the advent of increased awareness of terrorism, nerve agents had gained prominence as weapons of mass destruction. In these compounds, the OPs are a group of both synthetic and biogenic OP compounds, characterized by the presence of the binding covalent, carbon to phosphorus (C-P) bond. In OPs, this carbon to phosphorus bond replaces one of the four carbon-to-oxygen-to-phosphorus bonds of the more common phosphate ester (Wanner and Metcalf, 1992). While the vast majority of phosphorus-containing organic compounds contained the phosphate ester bond, both synthetic and naturally occurring phosphonates were still of importance (Blackburn, 1981). The direct C-P linkage is chemically and thermally inert, with the result that most of organophosphonate compounds are resistant to chemical hydrolysis, thermal decomposition and photolysis compared to analogous compounds containing the more reactive N-P, S-P or O-P linkages (Figure 1). The letter R represents either ethyl or methyl. Phosphonate contains an alkyl(R-) in place of one alkoxy group (RO-). X is called leaving group and is the principal metabolite for a specific identification (Kazemi et al., 2012).
Figure 1: General chemical composition of organophosphate pesticides
Source: Kazemi et al., (2012).
1.2.4 Pesticide metabolism
Presently, there are more of 900 pesticides and more of 600 active pesticide ingredients in the market (Hall et al., 2001). Millions of tons of pesticides are applied annually; however, less than 5% of these products are estimated to reach the target organisms, with the remainder being deposited on the soil and non-target organisms, as well as escaping into the atmosphere and water (Pimental and Levitan, 1986). The metabolic fate of pesticides is dependent on abiotic factors (temperature, moisture, soil pH, etc.), microbial community or plant species (or both), pesticide characteristics (hydrophilicity, pKa/b etc.), and biological and chemical reactions. Abiotic degradation is due to chemical and physical transformations of the pesticide by processes such as photolysis, hydrolysis, oxidation, reduction, and rearrangements. Further, pesticides may be biologically unavailable because of compartmentalization, which occurs as result of pesticide adsorption to soil and soil colloids without altering the chemical structure of the original molecule. However, enzymatic transformation, which is mainly the result of biotic processes mediated by plants and microorganisms, is by far the major route of detoxification. Metabolism of pesticides may involve a three-phase process (Shimabukuro, 1985; Hatzios, 1991) (Figure 2). In the first phase involving metabolism, the initial properties of a parent compound are transformed through oxidation, reduction, or hydrolysis to produce a more water-soluble and less toxic product than the parent. The second phase involves conjugation of a pesticide or pesticide metabolite to a sugar, amino acid, or glutathione, which increases further, the water solubility and reduces toxicity compared to the parent pesticide. Phase II metabolites were reported to have little or no phytotoxicity and may be stored in cellular organelles. The third phase involves conversion of Phase II metabolites into secondary conjugates, which are also nontoxic (Hatzios, 1991). In leafy spurge (Euphorbia esula L.), examples of Phase III metabolism are the conjugation of the N-glycoside metabolite of picloram with malonate and the formation of a gentibioside from the picloram glucose ester metabolite (Frear et al., 1989). McKellar et al. (1976) fed chlorpyrifos to dairy cattle for 2 weeks. The parent compound and two (oxidized and hydroxylated) metabolites were found at low levels in milk and cream (fat) and the concentrations of all three compounds decreased rapidly after cessation of administration. Hsu et al. (1995) recovered a maximum of 0.14% of intake via the eggs and depletion of residues from the body was rapid. Researchers had shown that pesticides have serious deleterious effects on the rumen fluid (Cook, 1969). Cook (1957) was perhaps the first to suggest that rumen liquor played an active role in hydrolyzing OPs, particularly parathion. Additional evidence by Cook (1957) indicated that metabolism of parathion by rumen microorganisms accounted for its apparent lack of toxicity to cattle. Certain OP pesticides were shown by Williams et al. (1963) to stimulate gas production In vitro by rumen holotrich protozoa, whereas these compounds had no appreciable effect when rumen bacteria served as the inoculums source.
Figure 2: Schematic metabolism of pesticides in body
Source: Barr and Needham (2002).
1.2.5 Mechanism of action of OP pesticides
OP pesticides avidly bind to acetyl cholinesterase (AChE) molecules and share a similar chemical structure (Figure 3). This leads to accumulation of acetylcholine and subsequent over-activation of cholinergic receptors at the neuromuscular junctions and in the autonomic and central nervous systems (Paudyal, 2008). The rate and degree of AChE inhibition differs according to the structure of the OP compounds and the nature of their metabolites. After the initial formation of AChE-OP complex, two reactions take place; at the first reaction, spontaneous reactivation of the enzyme may occur at a slow pace, much slower than the enzyme inhibition and requiring hours to days to occur. The rate of this regenerative process solely depends on the type of OP compound: spontaneous reactivation half life that lasts 0.7 h for dimethyl and 31 h for diethyl compounds. In general, AChE-dimethyl OP complex spontaneously reactivate in less than one day whereas AChE-diethyl OP complex may take several days and reinhibition of the newly activated enzyme can occur significantly in such situation. The spontaneous reactivation can be hastened by adding nucleophilic reagents like oximes, liberating more active enzymes. These agents thereby act as an antidote in OP poisoning (Eddleston et al., 2002). At the second step, with time, the AChE enzyme-OP complex loses one alkyl group making it no longer responsive to reactivating agents. This progressive time dependent process is known as ageing. The rate of ageing depends on various factors like pH, temperature, and type of OP compound; dimethyls OPs have ageing half life of 3.7 h whereas it is 33 h for diethyl OP (Worek et al., 1999). The slower the spontaneous reactivation, the greater the quantity of inactive AChE available for ageing. Oximes, by catalyzing the regeneration of active AChE from enzyme-OP complex, reduce the quantity of inactive AChE available for ageing. Since ageing occurs more rapidly with dimethyl OPs, oximes are hypothetically useful before 12 h in such poisoning. However, in diethyl OP intoxication, they may be useful for many days (Worek et al., 1999).
Figure 3: Schematic inhibitory of AChE by organophosphate pesticides
Source: Paudyal (2008)
1.2.6 Oxidative stress induced by pesticides
Reactive oxygen species (ROS), which include all highly reactive, oxygen-containing molecules, are an important part of the defense mechanism against infection, but excessive generation of ROS may damage tissues. ROS formed under both physiological and pathological conditions in mammalian tissues are capable of reacting with membrane lipids, nucleic acids, proteins, and enzymes and other small molecules, resulting in cellular damage. When balance between ROS and antioxidant system is lost “oxidative stress” results. ROS were found to be involved in infertility as evidenced by defective sperm function and in cryptorchidism upon exposure to toxic chemicals. It was shown that the cytochrome P450 enzymes of the steroidogenic pathway use molecular oxygen and electrons transfer from nicotinamide adenine dinucleotide phosphate (NADPH) to hydroxylate the substrate. In this process, superoxide anion or other oxygen free radicals were produced as a result of electron leakage in normal reactions or due to interaction of steroid products or other pseudosubstrates with enzymes. The antioxidant system plays an effective role in protecting testes and other biological tissues below a critical threshold of ROS thus, preventing testicular dysfunction. Antioxidant enzymes constitute a mutually supportive team of defense against ROS (Oschsendorf, 1999).
In living organisms, various ROS are generated in different ways, including normal aerobic respiration, stimulated polymorphonuclear leukocytes, and macrophages and per- . oxisomes. These appear to be the main endogenous sources of most of the oxidants produced by cells. Exogenous sources of free radicals include tobacco smoke, ionizing radiation, certain pollutants, organic solvents, and pesticides (Barlow et al., 2005). Pesticides have also been shown to initiate peroxidation of lipids in biological membranes (Koryagin et al., 2002; John et al., 2001). In erythrocytes, OP was found to produce morphological changes that are associated with increased LPO (John et al., 2001; Thapar et al., 2002). The effect of lipid peroxidation (LPO) on membrane lipids, membrane receptors, and membrane-bound enzymes disturbs function, structure, and fluidity of membranes and may result in altered ion flux (Halliwell and Chirico 1993). Generation of oxidative stress and consequent LPO by pesticides was reported in rat and human brain (Ranjbar et al., 2002).
Pesticides may irritate lung macrophages resulting in the generation of superoxide radical. Pesticides may be more reactive with an oxygen free radical that re-oxidizes to form superoxide. The pesticide may itself be a free radical, or may deplete antioxidant defenses. The overall effect is the production of more free radicals. The activities of superoxide dismutase (SOD), glutathione (GSH) peroxides, and glutathione reductase (GSH) are decreased owing to consumption of enzymes to neutralize free radicals generated by pesticides (Amer et al., 2002).
Earlier studies in animals also showed that pesticides such as paraquat 2,4-D, and endosulfan inhibit the activity of erythrocyte-SOD and induce oxidative stress in hepatocytes as well as in the central nervous system (Julka et al., 1993). Moreover, current research indicates that many widely used agricultural chemicals induce oxidative damage in various systems of the body, such as in dopaminergic cells of the brain by modulating the antioxidant defense system (Barlow et al., 2005). Increased lipid peroxide levels owing to exposure of OP and carbamate pesticides were reported in several studies (Dave, 1998; Prakasam and Sethupathy, 2001; Patil et al., 2003).
1.2.7 Organophosphate pesticdes, oxidative stress, and antioxidant status
OP insecticides, apart from inhibition of AChE and presence of cholinergic effects (Ren~o’n-von Osten et al., 2005; Bajgar et al., 2009), hyperglycemia (Joshi and Rajini, 2009), and oxidative stress were reported as adverse effects in poisoning by OP in both humans and animals. Oxidative stress induced by OP leads to disturbances in the function of different organs and tissues. In subchronic or chronic OP, exposure induction of oxidative stress was shown by various investigators as the predominant mechanism of toxicity (Giordano et al., 2007; Bűyűkokuroglu et al., 2008).
A case control study was conducted to evaluate the existence of oxidative stress, balance between total antioxidant capacity and oxygen free radicals in patients following acute OP exposure. The results showed significant LPO accompanied with decreased levels of total antioxidant capacity, total thiols, and AChE activity. A significant correlation existed between AChE suppression and reduced total antioxidant capacity (Ranjbar et al., 2002).
Fenthion (O,O-Dimethyl O-[3-methyl-4-(methylsulfanyl)phenyl] phosphorothioate) (Figure 4) is a contact and stomach insecticide used against many sucking and biting pests. It is particularly effective against fruit flies, leaf hoppers, cereal bugs, stem borers, mosquitoes, animal parasites, mites, aphids, codling moths, and weaver birds. It has been widely used in sugar cane, rice, field corn, beets, pome and stone fruit, citrus fruits, cotton, olives, coffee, cocoa, vegetables, and vines (Extension Toxicity Network, 2003). Based on its high toxicity on birds, fenthion has been used to control weaver birds and other pest-birds in many parts of the world. Fenthion is also used in cattle, swine, and dogs to control lice, fleas, ticks, flies, and other external parasites (United States Environmental Protection Agency, 2001; EXTOXNET, 2003; Australian Pesticides and Veterinary Medicines Authority, 2005).
Amid concerns of harmful effects on environment, especially birds, Food and Drug Administration no longer approves uses of fenthion. However, fenthion has been extensively used in Florida to control adult mosquitoes. After preliminary risk assessments on human health and environment in 1998 and its revision in 1999, USEPA issued an Interim Reregistration Eligibility Decision (IRED) for fenthion in January 2001. The EPA has classified fenthion as Restricted Use Pesticide (RUP), and warrants special handling because of its toxicity (Agency for Toxic Substances and Disease Registry, 2005).
Figure 4: Structure of fenthion
Source: Royal Society of Chemistry (2015).
Fenthion exposure to general population is quite limited based on its bioavailability. Common form of fenthion exposure is occupation related, and occurs through dermal contact or inhalation of dust and sprays (ATSDR, 2005). Another likely means of contamination is through ingestion of food, especially, if it has been applied quite recently with fenthion. So far, ingestion is the most likely severe poisoning case on humans and animals. To avoid this, crops applied with fenthion should be allowed enough degradation time before harvesting. Normally, 2 – 4 weeks time is enough depending upon the type of crop. It is a moderately toxic compound in EPA toxicity class II (probable human carcinogen). It is classified by the U.S Environmental Protection Agency (EPA) as a restricted use pesticide due to the special handling warranted by its toxicity. The half life of fenthion in water under field conditions has been reported to range from 2.9 to 21.1 days for various ocean, river, swamp, lake waters (Vlastos and Ganidi, 2004). Very few studies have demonstrated the effect of fenthion on fish (Thomas and Murthy, 1976).
1.2.9 Description of Clariidae
The family Clariidae belongs to the order Siluriformes and contributes significantly to annual freshwater fish production in South and South-East Asia and Africa (Na Nakorn, 1999). This family is naturally distributed all over Africa, South and South-East Asia with the highest genetic diversity found in Africa (Ilaboya, 2011). Nearly one fifth of all known catfish species occur in Africa and South-East Asia, however, the highest diversity is found in Africa with 14 genera and 92 species (Teugel, 1986), while only two genera with 17 species are presently known from Asia (Teugels, 1996).
Generally, Clariidae catfishes are elongated, have long dorsal and anal fins, and four pairs of barbels. A remarkable characteristic of this family is the possession of suprabranchial organ, formed by folds of the second and fourth branchial arches. With this organ, they are able to practice aerial respiration, implying that they can survive out of water for a long time. They are also known for walking on land over distances of several hundred meters, breathing atmospheric air and using their pectoral spines as a support (Teugels and Gourѐne, 1997).
The genus Clarias is the most common and popular of the Clariidae containing 32 species in Africa (Teugels, 1986), one of which is Clarias gariepinus (Burchell, 1822). It is of great economic importance as it is the most cultured catfish in Africa and the third most cultured catfish species in the world (Garibaldi, 1996). In Nigeria, it is the most abundant culturable fish apart from Tilapia (Elliot, 1985). Clarias gariepinus was selected for this study because it is of commercial importance as well as an aquaculture candidate that can narrow the gap between demand and supply of animal protein in developing countries (Nwani et al., 2013a). The species serve as an attractive model for toxicity testing due to its availability throughout the season, wide distribution in the environment being open swimmers, food source for human (Amin and Hashem, 2012) and ability to acclimate easily to laboratory conditions (Nwani et al., 2013a).
Research has shown that Cyprinus carpio exposed to sub lethal concentrations (0.38, 0.19, 0.96 mg/l) of fenthion suffered alterations in blood parameters. There was decrease in RBC count, hemoglobin concentration, ESR, and clothing time whereas WBCs count and platelet count were found to be increased (Muralidharan, 2012). The author concluded that fenthion is moderately toxic to C. carpio based on the 96 h LC50 recorded. Exposure to chronic sublethal concentrations of fenthion suggests that treated fish are faced with severe metabolic stress as a result of blood alterations thereby indicating that the use of this pesticide in the fields may be a threat to fishes as well as humans. Another study on the same species revealed that its exposure to fenthion destroyed melanophores of the exposed fish and the destructive changes observed is directly related to the strength of the dose induced (Muralidharan and Pillai, 2012). The result showed that melanophores changes noted is a significant abnormality providing a symptomatic index of toxicity.
Previous studies showed that fenthion caused significant disturbances in glutathione (GSH) redox status in the liver and brain of Cyprinus carpio (Sevgiler et al., 2007; Uner et al., 2009) and in the brain of Oreochromis niloticus. Research has also shown that fenthion affected the GSH redox cycle in a tissue-specific manner (Sevgiler and Uner, 2010). According to the authors, while GSH content decreased, GST activity increased in the liver and decreased in the kidney but the same concentration of fenthion did not affect the brain GSH contents and GST activity of the same fish. The studies above showed that fenthion altered the haematological parameters as well as antioxidant enzymes in humans, frogs, Cyprinus carpio but its effects on Clarias gariepinus is unknown hence, the need for the present study.
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