In this study, selected water quality parameters, concentrations of natural estrogenic steroids; estrone (E1), and 17β-estradiol (E2) and a synthetic steroid; 17α-ethynylestradiol (EE2) as well as total estrogens in 33 water samples from the Napoleon Gulf of Lake Victoria, Uganda and 3 comparative water samples from Lubigi channel, Kampala, in the month of May, 2013 were determined. The water samples were subjected to C18 solid phase extraction before analysis. In addition, biological endpoints such as gonadosomatic index (GSI), hepatosomatic index (HSI), and histopathological changes were assessed in an attempt to investigate the overall health and production performance of Oreochromis niloticus caught from Napoleon Gulf.
Lake Victoria is the world's second largest fresh water body in surface area covering 68800 km2 and is second only to Lake Superior in North America. The lake basin supports about 30 million people in the five countries of Kenya, Uganda, Tanzania, Burundi, and Rwanda and is the source of River Nile. The lake together with the River Nile is a resource of great socio-economic potential to over 160 million people living in various countries including Uganda, Tanzania, Egypt, Sudan, Ethiopia, Eritrea, Kenya, Rwanda, Democratic Republic of Congo and Burundi. However, a rapid increase in human population around the lake catchment area has had far-fetching impacts on the dynamics of the lake water resources. In the past few decades, increased discharges of poorly treated municipal wastes into the lake has been reported.
This study aims to test the water of Lake Victoria for possible pollution, draws conclusions and gives recommendations.
TABLE OF CONTENTS
DEDICATION
ACKNOWLEDGEMENT
ABBREVIATIONS AND ACRONYMS
LIST OF TABLES
LIST OF FIGURES
ABSTRACT
CHAPTER ONE
1.0 INTRODUCTION
1.1 Background
1.2 Problem statement
1.3 Study objecties
1.3.1 General objectie
1.3.2 Specific objecties
1.4 Research Questions
1.5 Justification and significance of the study
1.6 Scope of the study
CHAPTER TWO
2.0 LITERATURE REVIEW
2.1 Potential sources and effect of water pollutants in Lake Victoria
2.2 Sources of Lake Victoria water pollutants in Kenya and Tanzania
2.3 The physicochemical parameters of water quality
2.3.1 pH
2.3.2 Dissoled Oxygen
2.3.3 Temperature
2.3.4 Electrical conductiity/ionic strength
2.5 Occurrence and bioaailability of endocrine disrupting estrogens in enironment
2.6 Concentrations of estrogens in different enironmental matrices
2.7 Mechanisms of action of endocrine disrupting chemicals (EDCs) in -ertebrates
2.8 The pharmacokinetics of estrogens in animals and humans
2.9 Detection of EDEs in water matrix
2.10.1 Gonadosomatic index and hepatosomatic index
2.10.2 Histological techniques
2.11 Histopathological effects of EDEs in fish
2.11.1 Gills
2.11.2 Lier
2.11.3 Kidneys
2.11.4 Muscle
2.11.5 Testes and oaries
CHAPTER THREE
3.0 MATERIALS AND METHODS
3.1 Research design
3.2 Sampling techniques
3.3 Location of the sampling sites
3.4 Assessment of water quality at sampling points in Napoleon Gulf, Lake Victoria
3.5 Collection of water and fish samples for analysis
3.6 Isolation of estrogens from water samples
3.7 ELISA determination of estrogen concentrations in water samples
3.7.1 Determination of total estrogens
3.7.2 Determination of estrone
3.7.3 Determination of 17β-estradiol
3.7.4 Determination of 17α-ethynylestradiol
3.8 Calculation of gonadosomatic and hepatosomatic indices of O. niloticus
3.9 Preparation of fish tissue samples for histopathological analysis
3.9.1 Sectioning of fixed samples
3.9.2 Tissue processing and staining
3.9.3 Microscopic obseration of heamatoxylin and eosin stained tissues
3.10 Data recording
3.11 Data Analysis
3.12 Ethical issues
CHAPTER FOUR
4.0 RESULTS
4.1 Water quality Parameters
4.2 Concentration of total estrogens in water
4.3 Concentration of estrone in water
4.4 Concentration of 17β-estradiol in water
4.5 Concentration of 17α-ethynylestradiol in water
4.6 Variations in estrogen concentrations in water samples from Napoleon Gulf
4.7 Endpoints in fish health and production
4.7.1 The gonadosomatic index (GSI)
4.7.3 The hepatosomatic index (HSI)
4.8 Histopathological analysis
CHAPTER FIVE
5.0 DISCUSSION OF RESULTS
5.1 Physicochemical parameters of water quality
5.2 Total estrogens in water samples
5.3 Estrone in water samples
5.4 17β-estradiol in water samples
5.5 17α-ethynylestradiol in water samples
5.6 Estrogens in waste water samples from Lubigi channel, Kampala
5.7 Fish health and Production indices
5.8 Histopathological findings
CHAPTER SIX
6.0 CONCLUSIONS AND RECOMMENDATIONS
6.1 Conclusions
6.2 Recommendations
REFERENCES
APPENDICES
APPENDIX I: ELISA absorbance readings of estrogens in the sample, standards and controls using a spectrophotometer (Gene 5) at 450 nm
APPENDIX II: Measurement and collection of organs from fish in pictures
APPENDIX III: Different wastewater sample collection sites in pictures
APPENDIX IV: Collection of fish samples and laboratory analysis of estrogens in water samples at Napoleon Gulf of Lake Victoria and Makerere Uniersity respectiely in pictures
APPENDIX V: Standard cures for ELISAs
Figure 5C & 5D: The 17α-ethynylestradiol ELISA has a detection range between 0.05 and 3.0 µg/L (Figure 5C). The ELISA for total estrogens has a detection range of 0.05 and 3.0 µg/L (Figure 5D)
APPENDIX VI: Water sampling sites in Napoleon Gulf of Lake Victoria and some areas in Kampala Uganda
APPENDIX VII: Some of the histopathological changes obsered in different tissues of O. niloticus caught from Napoleon Gulf of Lake Victoria, Uganda
APPENDIX VII: Data collection forms
DEDICATION
I dedicate this research work to my dear wife Ms. Kyarikunda Rose, beloed daughter Ekyarikunda Menten, beloed mum Gertrude Bamwenzaki, brothers; Mushabe Naboth and Kabesiime Frank & sisters; Namurinda Allen, Twongyeirwe E-ace, Natukunda Rosette and Nagasha Emily for their continued support and encouragement towards my academic endeaors.
ACKNOWLEDGEMENT
My sincere thanks go to Assoc. Prof. Byarugaba Denis for his tireless guidance to me from the -ery beginning up to completion of this research work. May the almighty God reward you accordingly!
I am also grateful to Dr. Vuzi Peter for cosuper-ision of this work, continuous encouragement and inspirations especially during difficult moments of my research career.
I would like to acknowledge the technical guidance and adice of Dr. Okuni Boniface in describing the histopathological data. He has made the histopathology work easier and interesting.
I also appreciate the efforts of Mr. Kizito Muwonge in proiding me with all the equipment used in the measurement of water quality characteristics.
My special thanks also go to Assoc. Prof. Owiny Daid, Drs; Ikwap Kokas and Mugizi Dennis for permission to use their Laboratory space and equipment during extraction of estrogens from water samples.
My sincere gratitude also goes to Dr. Vudriko Patrick for technical adice in the analysis of ELISA data.
Finally, I acknowledge the contribution of Birungi Doreen, Kayaga Beatrice, Asiimwe Rwatooro, Kato Gerald, Sabiiti Eria, Abuo Martha, Ibanda I-an, Nalunga Justine, Nturanemigisha Herbert, and Musingwire William in extraction of the water samples for analysis of endocrine disrupting estrogens.
This study was made possible through funding from The Inter-Uniersity Council of East Africa’s VicRes Grant No. Vic/P14/07 awarded to Assoc. Prof. Denis Byarugaba.
ABBREVIATIONS AND ACRONYMS
Abbildung in dieser Leseprobe nicht enthalten
LIST OF TABLES
Table 1: The estrogenic endocrine disruptors pathways to waterways
Table 2: Reported estrogen concentrations in different enironmental matrices in New Zealand
Table 3: Sampling site description in the study areas
Table 4: The mean -alues of physicalchemical parameters of water at different sampling points in Napoleon Gulf of Lake Victoria, Uganda
Table 5: Concentrations of total estrogens (ng/L) in water samples from different sampling points in Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda
Table 6: Concentrations of estrone (ng/L) in water samples from different sampling points in Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda
Table 7: Concentrations of 17β-estradiol (ng/L) in water samples from different sampling points in Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Table 8: Concentrations of 17α-ethynylestradiol (ng/L) in water samples from different sampling points in Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Table 9: Mean GSI (%) and HSI (%) of female and male fish from Napoleon Gulf of Lake Victoria, Uganda
LIST OF FIGURES
Figure 1: Mechanisms of action of endocrine disruption
Figure 2: Location of the Napoleon Gulf of Lake Victoria showing the study areas
Figure 3: Flow chart of Solid Phase Extraction (SPE) of estrogens from water sampl
Figure 4: Mean -alues of water quality parameters at different sampling points in Napoleon Gulf of Lake Victoria, Uganda
Figure 5: Concentrations of total estrogens (ng/L) in water samples from different sampling points in Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Figure 6: Concentrations of estrone (ng/L) in water samples from Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Figure 7: Concentrations of 17β-estradiol (ng/L) in water samples from Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Figure 8: Concentrations of 17α-ethynylestradiol (ng/L) in water samples from Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Figure 9: Variations in the concentration of total estrogens, 17β-estradiol, 17α-ethynylestradiol, estrone in water samples from Napoleon Gulf of Lake Victoria, Ugand
Figure 10: Mean GSI (%) of fish caught from Napoleon Gulf of Lake Victoria, Ugand
Figure 11: Mean HSI (%) of fish caught at different sites in the Napoleon Gulf of Lake Victoria, Uganda
Figure 12: Spleen tissue of O. niloticus (X100) stained with hematoxylin and eosin showing siderotic nodules
Figure 13: Lier tissue of O. niloticus (X100) stained with hematoxylin and eosin showing leukocyte infiltrations
Figure 14: Testicular tissue of O. niloticus (x 100) stained with hematoxylin and eosin showing extensie areas of necrosis and fibrosis
Figure 15: O-arian tissue of O. niloticus (x 100) O. niloticus (X100) stained with hematoxylin and eosin showing different deelopmental stages with associated histopathologies
Figure 16: O-arian tissue of O. niloticus (x 100) O. niloticus (X100) stained with hematoxylin and eosin showing different deelopmental stages with associated peculiar histopathologies
ABSTRACT
The occurrence of steroid hormones in different enironmental matrices including water has a public health implication because of their strong endocrine disrupting potency. In the present study, selected water quality parameters, concentrations of natural estrogenic steroids; estrone (E1), and 17β-estradiol (E2) and a synthetic steroid; 17α-ethynylestradiol (EE2) as well as total estrogens in 33 water samples from the Napoleon Gulf of Lake Victoria, Uganda and 3 comparatie water samples from Lubigi channel- Kampala in the month of May, 2013 were determined. The water samples were subjected to C18 solid phase extraction before analysis. In addition, biological endpoints such as gonadosomatic index (GSI), hepatosomatic index (HSI), and histopathological changes were assessed in an attempt to inestigate the oerall health and production performance of Oreochromis niloticus caught from Napoleon Gulf. Results from water quality analysis were suggestie of lake water pollution. The detection ranges of estrogens in water samples from Napoleon Gulf were; 200 and 800 ng/L; 9.8 and 49 ng/L; 145 ng/L and 305 ng/L; 45 ng/L and 360 ng/L for total estrogens, estrone, 17β-estradiol, and 17α- ethynylestradiol respectiely. The concentrations of estrogens in water samples from Kampala were higher than those detected in water samples from Napoleon Gulf except for 17α-ethynylestradiol (p < 0.05). Multiple comparisons of estrogen concentrations in the three categories of water samples showed significant -ariations. Conersely, the histopathological changes in different tissues were quite similar and included focal areas of necrosis, macrophages infiltrations, siderotic nodules, fibrosis, and melanophages in the spleen, lier, and testis as well as atretic follicles in oarian tissue. The obsered histopathological changes could hae resulted from exposure of fish to -arious waterborne chemical pollutants indicating water pollution. The presence of higher concentrations of estrogens in wastewater samples from Kampala compared to samples from Napoleon Gulf could be suggestie of the probable source of steroid hormones. Generally, analysis of water samples for estrogens reealed higher concentrations (about 9 times the lowdose effect; 1 ng/L) than those of similar preious studies. Equally, the histopathological changes obsered in the different fish tissues allude to the possible effects of -arious water chemical pollutants including endocrine disrupting estrogens (EDEs). Thus, the bioaccumulation potentials of steroid hormones in -arious tissues, indiidual and combinational effects of EDEs in fish and humans need further inestigations.
CHAPTER ONE
1.0 INTRODUCTION
1.1 Background
Lake Victoria is the world’s second largest fresh water body in surface area coering 68800 km2 and is second only to Lake Superior in North America. The lake basin supports about 30 million people in the fie countries of Kenya, Uganda, Tanzania, Burundi, and Rwanda and is the source of Rier Nile. The lake together with the Rier Nile is a resource of great socioeconomic potential to oer 160 million people liing in -arious countries including Uganda, Tanzania, Egypt, Sudan, Ethiopia, Eritrea, Kenya, Rwanda, Democratic Republic of Congo and Burundi Odada et al. (2004). It is estimated that the growth rate in human population within 100 km of the Lake Victoria catchment area is 7%, a reflection of a growing dependence and pressure on the lake’s resources (Matagi, 2002). Howeer, the rapid increase in human population around the lake catchment area has had farfetching impacts on the dynamics of the lake water resources. In the past few decades, increased discharges of poorly treated municipal wastes into the lake has been reported (Kishe, 2004).
Recent studies hae found that endocrine disrupters are among the most common contaminants of -arious enironmental matrices including water bodies (Liu et al., 2011; Manickum et al., 2011; Xu et al., 2012). These are foreign substances or mixtures that alter function(s) of the endocrine system consequently harming an indiidual life form, its offspring, or population (Pothitou & Voutsa, 2008). E-en though there are seeral endocrine disrupting chemicals such as Bisphenol A (BPA), Nonylphenol (NP), Nonylphenol Ethoxylates (NPnEO) and Octyphenol, the principal actors responsible for the estrogenic actiity in water effluents are steroidal endocrine disrupters including the natural; estrone (E1), 17β-estradiol (E2), estriol (E3) and the synthetic ones; 17α- ethynylestradiol (EE2) present in birthcontrol pills (-on Saal et al., 2012).
Despite haing seeral sources of endocrine disrupting estrogens (EDEs), effluents from wastewater treatment plants are considered a major source of these substances in the aquatic enironment because are not totally remoed or degraded by physical, chemical, or biological treatment processes (Liu et al., 2011). It has been reported that humans excrete hormones both natural and synthetic estrogens from their bodies into the enironment through sewage discharge (Swart & Pool, 2007). The steroid hormones are excreted either in the free form or as conjugates into the enironment. Unlike peptide hormones which are readily destroyed in the enironment, steroid hormones are enironmentally stable (Huang et al., 2013). Other potential sources of EDEs into surface and ground waters are discharges from industrial manufacturing processes, leaks from sewer pipes, sewage oerflows, onsite wastewater systems, and animal feeding operations (Hongo & Masikini, 2003). Therefore, determination of the concentration of estrogens in different wastewaters in Uganda would proide releant information on probable sources of these substances to the -arious stakeholders inol-ed in the management of water resources.
The detection of estrogens in water has become an area of intensie research since these hormones hae been found to induce seere effects on reproductie performance in wildlife and humans (Cakmak et al., 2006). The presence of oa-testis in males and atretic oocytes in female fish has been reported as the main effects of EDES (Leino et al., 2005; Liu et al., 2011). In humans, deelopment of breast and testicular cancer as well as tubal and uterine dysmorphologies, hae been reported as the major health effects of estrogens (Muhamed, 2008).
Fish which feed on smaller aquatic life, are likely to bioaccumulate water contaminants in higher quantities than other aquatic smaller life. The effect of this is that een extremely low concentrations of estrogens such as 110 ng/L of 17β-estradiol, or as low as 0.1 ng/L in case of 17α-ethynylestradiol, can cause reproductie failure in fish (Young et al., 2002). Although seeral studies hae been done to detect leels of estrogens in different enironmental matrices including surface waters, no study has been done in Uganda to detect the presence of EDEs in Lake Victoria water and assess their potential effects on aquatic wildlife.
1.2 Problem statement
The presence of estrogenic compounds in water has been noted since the early 1980s (Gabet et al., 2007) and has become a great concern because they interfere with the reproduction of man, liestock, and aquatic life (Peng et al., 2008). In Uganda, Lake Victoria which is the main source of water and fish is also a sink of industrial and domestic effluents. These effluents are potential sources of estrogenic endocrine disrupters particularly steroid hormones of natural and synthetic origin that are excreted by humans and animals in large quantities. It has been found that prolonged exposure of fish and humans to these estrogens and estrogenic analogues causes changes in physiological and biochemical processes by interfering with the hormonal control of these life processes (Adam et al., 1993). Although seeral reports on the leels of endocrine disrupting estrogens in water and histopathological alterations in fish tissues exist, there is little data aailable on the concentrations of these substances in Lake Victoria water, Uganda and histopathological changes in tissues of Oreochromis niloticus from the same area. Therefore, the aim of the present study was to analyze the concentrations of estrogens in water samples from Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda. Biological endpoints such as gonadosomatic, hepatosomatic indices and histopathological changes were also analyzed to assess the effect of waterborne chemical pollutants including EDEs on reproductie performance of O. niloticus.
1.3 Study objecties
1.3.1 General objectie
To assess the leels of endocrine disrupting estrogens in water and study the pathological changes in tissues of O.niloticus from Napoleon Gulf of Lake Victoria, Uganda.
1.3.2 Specific objecties
The specific objecties of this study were to determine;
1. The physicochemical characteristics of water at different sampling points in Napoleon Gulf of Lake Victoria, Uganda.
2. The concentration of estrogens in water samples from Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda.
3. The gonadal and hepatosomatic indices as well as identify and describe pathological alterations in tissues of O. niloticus from the same enironment.
1.4 Research Questions
1. What are the physicochemical characteristics of water at different sampling points in Napoleon Gulf of Lake Victoria, Uganda?
2. What are the concentrations of estrogens in water samples from Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda?
3. What are the gonadal and hepatosomatic indices as well as obserable pathological alterations in tissues of O. niloticus from Napoleon Gulf of Lake Victoria, Uganda?
1.5 Justification and significance of the study
This study has added on the information aailable on pollution of Lake Victoria since none of the preious studies (Matagi, 2002; Naigaga et al., 2011; Odada et al., 2004; Ssebugere et al., 2013) focused on estrogen leels. It has also proided information on the histopathological changes in fish from the polluted water sources. Besides, the data generated will be of -alue to the -arious stakeholders inol-ed in enironmental protection and public health. Particularly, this information will be useful in strengthening appropriate measures required in the minimization of lake pollution and control of human health hazards. In addition, analysis of EDEs in the lake water has proided useful information required in monitoring of discharge sources and assessing the risk/dangers posed by estrogens to the aquatic wildlife and humans. Since fish act as biomonitors of water pollution and share many biological links with humans, the presence of higher concentrations of endocrine disrupting estrogens in water and the histopathological changes obsered in -arious tissues of O. niloticus will make this fish species act as an experimental model for endocrine disrupting estrogens testing on waters of Lake Victoria.
1.6 Scope of the study
This study was field and laboratory based inol-ing collection of water and fish samples from the field and their analysis in the laboratory. It was limited to only eleen sampling stations and three comparatie sampling points in Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda respectiely. Water samples were collected from these areas and used in the determination of the concentration of endocrine disrupting estrogens using competitie ELISA techniques. Concurrently, testis, lier, spleen and oary tissues from fresh samples of O. niloticus obtained from and/ or near study areas were used in examination of histopathological changes. In situ measurements of pH, dissoled oxygen (DO), electrical conductiity (EC), and temperature were also undertaken to assess the water quality in the study area.
CHAPTER TWO
2.0 LITERATURE REVIEW
2.1 Potential sources and effect of water pollutants in Lake Victoria
Irrespectie of the source, water pollution is caused by the actions of the people globally. Thus, the main sources of water pollutants are effluents from industries, septic tanks, waste disposal sites, sewage treatment plants among others. This has been exacerbated by the increasing human settlements, growth of industries and urbanization around different water bodies which has resulted into large discharges of industrial and municipal wastewaters into the lakes (Andersen et al., 2003).
The deficient sewage and industrial wastewater treatment plants, smallscale workshops, waste oil from parking lots and car repair garages are major sources of pollution load for Lake Victoria. Worse still, the sewer system in Kampala city seres only a small fraction of the city population and only 10% of all sewage generated in Kampala gets treated. Guesthouses, slum dwellings and industries discharge untreated wastewater directly into the Lake (Scheren et al., 2000). According to Matagi (2002), it was reported that the major causes of enironmental degradation in Kampala, the capital city of Uganda include poor solid waste collection, inadequate facilities for sewage and sanitation, poor drainage, increasing traffic and industrial pollution as well as urban agriculture. Unfortunately, most of the pollutants generated from these anthropogenic actiities in the city end up in Lake Victoria through channels draining from the city such as Nakiubo channel.
Jinja Municipality is one of the municipalities that are experiencing rapid growth and its nearby towns including Njeru, Buwenda, Kimaka, Mpumudde, Masese, Walukuba, Wanyange, Kakira, Wairaka and Bugungu. Various human actiities such as direct discharge of human waterwastes and improper disposal of left oer medications containing steroid hormones into the drainage channels exist in the Municipality and its neighboring towns. Subsequently, the domestic waterwastes hae been discharged into the lake through different drainage channels such Wanyange and Kakira in the Napoleon Gulf (Tegule, 2011). In addition, most wetlands in the area hae been turned into waste water and garbage discharge points (Scheren et al., 2000). For example Kirinya wetland located along the Northern shoreline of Lake Victoria in Jinja Municipality receies discharges of domestic, municipal and industrial effluent, surface runoff and storm water from Jinja town. Although this wetland is not an official waste dumping site, open waste dumping of both biodegradable and nonbiodegradable materials is practiced. Equally, Masese wetland receies urban effluents and storm water runoff through a 1.4 Km long drainage channel (Naigaga et al., 2011).
The Kenya’s side of the lake has not been spared either, towns of Kakamega and Kisumu discharge inadequately treated sewage in riers draining into Lake Victoria because of deficient treatment plants. Kisumu’s sewage plant at Kisat with a design capacity of 9000 cubic meters now receies 15000 cubic meters of influents, much of which flows into Lake Victoria without treatment. In addition, in Kakamega, sewage lagoons at Nabongo hae been neglected for more than 3 years and are grossly inefficient. Conersely, Tanzania’s town of Mwanza located near Lake Victoria discharges large quantities of untreated waste into the lake. More still, waste from fish processing factories, oil processing plants, textile facilities and tanneries located in the Mwanza area is discharged into the lake without pretreatment (Scheren et al., 2000).
Of great concern, these wastes contain biological and chemical pollutants that cause deleterious effects in humans and fish.
Human diseases such as typhoid, cholera, hepatitis, enteritis are caused by water borne pathogens. It was reported that water related diseases caused by insufficient safe water supplies coupled with poor sanitation and hygiene caused 3.4 million deaths per year particularly among children (Peng et al., 2008). On other hand, it has been reported that heay metals such as lead, mercury, selenium, arsenic cause cancer, mental retardation, miscarriages in humans. In fish, chemical contaminants hae caused necrosis of lier, kidneys, gills, spleen as well as the gonads (Andersen et al., 2003). In Uganda, it was reported that significant declines in fish stocks and fish species diersity in Lake Victoria were due to oerexploitation, use of destructie fishing gear and methods (Naigaga et al., 2011). The use of fishing nets of smaller sizes led to the trapping of immature fish thereby affecting the fish population and breeding patterns in the lake.
Despite -arious interentions such as burning of illegal fish nets and establishment of beach management units, declines in fish populations in Lake Victoria are still being registered. It has been reported that direct discharge of industrial and municipal waterwastes into the lake has greatly affected the ecosystem of the lake through pollution. Thus, inorganic and organic water contaminants such as lead, mercury, cadmium, tin; dioxin, PCB and DDT present in the waterwastes may be responsible for the fish declines (Naigaga et al., 2011). Furthermore, it has been reported that nearly half of the Lake Victoria floor experiences anoxia spells for seeral months of the year compared to the four decades ago when anoxia was sporadic and localized. In addition, algal biomass concentration is almost fie times greater in the surface waters today than reported in the 1960s (Odada et al., 2004). Scheren et al. (2000) reported that discharge of raw sewage, dumping of domestic and industrial wastes, fertilizers, and chemicals into the lake are the main actiities contributing to Lake Victoria pollution. This has been attributed to the presence of mushrooming industries, and floricultural gardens in and around the major towns and cities that include Kampala, Jinja, and Entebbe and lack of clear policies and guidelines on waste management.
2.2 Sources of Lake Victoria water pollutants in Kenya and Tanzania
In East African region, Lake Victoria is increasingly experiencing pollution from -arious sources like biomass burning, industry, transport and households (Ssebugere et al., 2013). Due to the eer increasing human population size, wetlands on the Ugandan side of Lake Victoria hae been extensiely degraded by human actiities such as habitat modification for agriculture, urbanizations and infrastructure proision, municipal and animal waste disposal. Furthermore, the establishment of smallscale economic enterprises such as informal slaughter houses, poultry units, motor garages and car washing facilities has greatly contributed to pollution of the lake. The causes of rising pollution leels in Lake Victoria are as many as they are dierse and each of the three East African nations is culpable. The Lake has for a long time been a sink to excessie nutrients and untreated effluents that hae led to fish dieoffs, algal blooms and the spread of hyacinth, a ferocious waterweed (Scheren et al., 2000).
2.3 The physicochemical parameters of water quality
Safe water is a precondition for health and deelopment and a basic human right. Thus, water quality is an important natural resource concern for the nations across the globe (Huber & Mosler, 2012). Water quality is simply and intuitiely defined as a term that describes the physical, chemical, and biological characteristics of water usually with respect to its suitability for a particular purpose (Kishe, 2004). Physicochemical parameters of water are good preliminary indicators of water quality used in ealuating the extent of water pollution. Therefore, it is important to measure these -ariables wheneer conducting water related studies. The following physicochemical parameters were determined in this study.
2.3.1 pH
This is the measure of the leel of hydrogen ions in solution that results in its acidic or basic quality. It is worth noting that each aquatic organism is adapted to a specific pH range, howeer, a pH of 7.4 is optimum for most fish species (Cleoni dos Santos et al., 2012). The pH in most water bodies unaffected by humans ranges from 6.5 to 8 (Kishe, 2004). Humans contribute to eleated pH primarily in the form of nutrient runoff containing fertilizers which leads to increased algal growth and a higher pH. Besides, low pH results when atmospheric oxygen and water come in contact with sulfides that react to form acids. In particular, O. niloticus can suri-e in a wide range of water acidity that ranges between pH of 3.5 and 12. Howeer, the recommended pH for good growth of Nile tilapia is 6.5 to 9 (Birungi et al., 2007). In addition, pH of water affects the solubilities of estrogens where at higher pH -alues (7-10) these compounds are significantly ionized and therefore more soluble in water. Howeer, the relatie increase in solubility at high pH is greatest for the leastsoluble compounds, E1 and E2 (Shareef et al., 2006). In the present study, pH was measured to appreciate its probable effects on fish growth and estrogen solubilities in water.
2.3.2 Dissoled Oxygen
In water bodies, the dissoled oxygen (DO) leels determine the ability of the water to support aquatic oxygen dependent life. Cold water holds more dissoled oxygen than warmer water and therefore season and time greatly influence the DO -alues. The DO measurements can be expressed as milliolts (mV) or milligrams per liter (mg/L) (Kishe, 2004). The DO limit for O. niloticus has been documented to be 0.01 mg/L. Howeer, the recommended DO leel for good growth of fish should ideally be aboe 2 mg/L (Birungi et al., 2007). The amount of DO is affected by the degree of eutrophication occurring in the lake and is dependent on organic load and in turn indicates the leel of lake pollution. In situations where DO decreases significantly, the ability of oxygen dependent organisms to suri-e is greatly reduced.
2.3.3 Temperature
In the assessment of water quality, temperature is one of the water characteristics measured since it affects nearly all other water quality parameters. It is worthy of note that different aquatic organisms are adapted to certain temperature ranges. The water temperature is regulated by solar energy and the surface area of the water body (Kishe, 2004). In particular, the optimal temperature for growth of most Nile tilapia species ranges from 25 to 28oC. Howeer, reproduction of Nile tilapia stops at 22oC and feeding below 20oC. In addition, tilapia cannot suri-e in temperatures less than 10 to 12oC for more than a few days (Hassan et al., 2013). Thus, measurement of water temperature in the study area was useful in the elucidation of the probable effects of temperature on the reproductie performance of O. niloticus.
2.3.4 Electrical conductiity/ionic strength
The ability of a substance to conduct an electrical current defines the term electrical conductiity and is measured in microsiemens per centimeter (mS/cm). Ions such as sodium, potassium, and chloride gie water its ability to conduct electricity. Conductiity is an indicator of the amount of dissoled salts in a water body. It is often an estimate of the amount of total dissoled solids rather than measuring each dissoled constituents separately (Kishe, 2004). Ionic strength has profound effect on the solubility of estrogens in water. The solubilities of estrogens decrease as the ionic strength increases due to the salting out effect. At higher ionic strength, estrogenic hormones aggregate and flocculate thereby reducing their solubility in water (Shareef et al., 2006).
2.4 Estrogenic endocrine disruptors (e-EDs) as water pollutants
In particular, estrogenic endocrine disrupters are exogenous agents that interfere with synthesis, secretion, transport, metabolism, binding action or elimination of natural bloodborne hormones that are present in the body and are responsible for homeostasis, reproduction, and deelopmental processes (Bourguignon & Parent, 2010). It has been reported that e-EDs do not only interfere with the endocrine system but also cause remarkable toxicological effects in aquatic and terrestrial animals (Mortensen et al., 2006). Thus, establishing potential sources and analysis of estrogens in different enironmental matrices in Uganda require urgent attention.
The sources of EDEs -ary depending on geographical location, but include both natural and anthropogenic origins. Natural sources include biologically deried estrogens excreted by humans, wild animals, and liestock. Plants and fungi produce estrogenic compounds. Animal deried estrogens are mainly introduced into aquatic systems through municipal wastewater systems and husbandry run off. Thus, the presence of numerous endocrine disrupting compounds (EDCs) in surface waters and sediment has been attributed primarily to their incomplete remoal in the surfacetreatment processes (Gomes et al., 2003). Therefore, determination of these chemicals in water is required to assess their enironmental impact. In addition, phytoestrogens present during algal blooms fueled by eutrophication are worthy of mention (Yong et al., 2014). Besides, combined sewage oerflows, onsite wastewater systems, and combined animal feeding operations are potential sources of e-EDs. Furthermore, the e-EDs are ubiquitous in our enironment and can be found in all media (water, air, soil); they are present in food products (soybeans, legumes, yams), and plants as phytoestrogens present in fruits, -egetables, beans and grasses. Estrogenic endocrine disruptors are also present in household products (detergents and associated surfactants), pesticides (DDT, endosulphan, atrazine, nitrofen) and plastics (bisphenol A, phthalates). Pharmaceuticals (birth control pills, cimetidine), industrial chemicals (PCBs, dioxin) and metals (cadmium, lead, mercury) are other examples of e-EDs (Jobling et al., 2003). It is therefore anticipated that analysis of estrogens in water from Lake Victoria, Uganda would proide reliable information on the leel of water contamination with these substances.
Table 1: The estrogenic endocrine disruptors’ pathways to waterways
Abbildung in dieser Leseprobe nicht enthalten
Adapted from Huang et al. (2013)
2.5 Occurrence and bioaailability of endocrine disrupting estrogens in enironment
Estrogens are excreted by humans in urine in conjugated forms with either a glucuronide or sulphate group linked to a hydrogen atom in the hydroxyl group of the steroid ring. Strikingly, the conjugated estrogens hae low estrogenic potency compared with the parent compounds. Howeer during their transfer from the wastewater treatment plants to surface waters, the conjugates are more or less hydrolyzed in their free form. Thus, significant proportions of conjugated hormones hae been reported to be present in surface waters (Gabet et al., 2007). Therefore, to assess accurately the concentrations of estrogens in aquatic enironment, it is essential to measure the total quantity of estrogens, including the conjugated fractions. Thus, in this study, the concentration of total estrogens was measured.
Conersely, the bioaailability, distribution, and persistence of organic compounds including estrogens in aquatic enironment can be partly influenced by their solubilities in aqueous media. The order of aqueous solubilities is associated with increasing polarities of the estrogens from estrone (E1-one hydroxyl group) to 17β-estradiol (E2-two hydroxyl groups) and then 17α-ethynylestradiol (EE2) with added ethynyl group at the 17α position of the sterol ring (Shareef et al., 2006). In the present study, the concentration of indiidual estrogens in water samples was determined.
2.6 Concentrations of estrogens in different enironmental matrices
The concentration of estrogens in different enironments -aries from country to country and from one enironmental matrix to another. The concentration of estradiol in sewage effluent was found to range from 2.7 to 48 ng/L in United Kingdom; 3ng/L in Germany; 64 ng/L in Canada and 55 ng/L in Japan. In Britain, the concentration of estrone in Sewage treatment work effluents -aried widely with a detection range of 1.4 to 76 ng/L whereas in Italy, the detection leels ranged from 2.5 to 8.21 ng/L. In Germany, a relatiely higher concentration of 70 ng/L of estrone was detected against the lower concentration of 48 ng/L in Canada. In parallel, it was reported that estriol was detected in the range of 0.43 to 18 ng/L in Italian sewage treatment effluent (Swart & Pool, 2007). In addition, Ying et al. (2009) reported that the aerage concentrations of estrogenic steroids; E3, E2, E1, and EE2 in influents of six Italian sewage treatment plants were 80, 12, 52, and 3 ng/L respectiely. Conersely, the concentrations of estrogenic steroids in the effluents ranged from below detection limit (< LOD) to 64 ng/L for E2, from less than LOD to 82 ng/L for E1, from 0.43 to 18 ng/L for E3 and from less than LOD to 42 ng/L for EE2. In South Africa, the mean score of 37 ng/L of estradiol was detected in influents compared to a relatiely lower mean leel of 5 ng/L of estradiol in effluent (Manickum et al., 2011). Estradiol concentrations of up to 7µg/L were detected in sewage effluents in the UK (Routledge et al., 1998). Furthermore, estradiol was detected at concentrations ranging from 1 ng/L to 80 ng/L in effluents from USA sewage treatment works (Wright-Walter & Volz, 2009). In contrast, the concentration of estrogens in different enironmental matrices in New Zealand -aried greatly (Jobling et al., 2003; Singhal et al., 2009) as shown in table 2. Thus, estrogens could be present in waters of Lake Victoria in concentrations that cause aderse health effects and thus negatiely impacting on the aquatic wildlife and humans consuming contaminated water.
Table 2: Reported estrogen concentrations in different enironmental matrices in New Zealand
Abbildung in dieser Leseprobe nicht enthalten
Adapted from Singhal et al. (2009)
2.7 Mechanisms of action of endocrine disrupting chemicals (EDCs) in -ertebrates
Although the mechanisms of action of EDC are poorly understood, EDCs interfere with the normal hormonal processes and regulation in one of the two ways; agonistic or estrogenic substances that bind to hepatic estrogen receptors mimicking natural endogenous estrogens. These hae effects of feminizing male fish or altering the normal hormonal control in females. Alternatiely, an agonist may compete with 17β-estradiol for the pituitaryhypothalamic feedback receptors that regulate egg deelopment. Secondly, antagonistic or antiestrogenic substances such as phenylethylenes may block hepatic estrogen receptor sites thereby preenting the normal interaction of estradiol (Prado et al., 2011). In addition, other interactions may occur affecting the synthesis and metabolism of hormones and alteration of hormone receptor leels in the exposed organism (Soto et al., 1995). Since the endocrine system is tuned to respond to -ery low concentrations, e-DCs act at extremely low serum concentrations, typically in the picomolar to nanomolar range (Pothitou & Voutsa, 2008).
Abbildung in dieser Leseprobe nicht enthalten
Adapted from Cooke et al. (2013)
Figure 1: Mechanisms of action of endocrine disruption
During reproductie steroidogenesis, P450 aromatase (CPY19) is responsible for the conersion of testosterone to estradiol. In fish, two forms of CYP19 exist that is CYP19A1 and CYP19A2 and are mainly present in the gonads and the brain respectiely although both are found in other tissues as well (Piferrer, 2011). The production of E2 is regulated at least in part by regulating CYP19 expression at transcriptional leel. Thus, stimulation of CYP19 by exogenous compounds leads to -arying results in fish depending on the isoform affected and the stage of deelopment. These effects range from sex reersal and impaired sexual differentiation to production of -itellogenin in males (Young et al., 2002). In particular, the synthetic estrogen 17α-ethynylestradiol stimulates CYP 19A2 expression but inhibits CYP19A1. Since CYP19A1 metabolizes estradiol and testosterone, its modulation by xenoestrogens are thought to affect circulating leels of actie hormones (Mortensen et al., 2006).
2.8 The pharmacokinetics of estrogens in animals and humans
The three main circulating forms of hormones in animal and human bodies are estrone (E1), 17β-estradiol (E2), and estriol (E3). They are rapidly absorbed from the gastrointestinal (GI) tract, skin, and mucous membranes, and after parenteral injection. Howeer, the unconjugated natural estrogens are rapidly inactiated in the GI tract and lier if taken orally. Thus, their deliery by other routes namely; transdermally, nasally or intramuscularly is warranted if they are to impart any significant aderse health effect. Micronized estradiols, steroidal estrogens that contain an ethinyl group at C17, the conjugated estrogens, and nonsteroidal estrogens are orally actie (Minneman et al., 2005). Once absorbed, they are frequently oxidized, hydroxylated, deoxylated and methylated in the lier prior to the final conjugation with glucuronic acid or sulfate (Ying et al., 2009). Howeer, estrogen sulfates are slowly excreted from the body and are therefore usually hydrolyzed in tissues to act as biologically actie estrogens (Minneman et al., 2005).
The pathway of estrogen metabolism starts with the conersion of estradiol to estrone. In turn, estrone is ultimately conerted to estriol or catechol estrogens that are then conjugated primarily with glucuronides, sulfates and thioesters. Thus, estrogens are excreted primarily as polyhydroxylated forms that are conjugated at C3 with sulfate or glucuronic acid. The conjugated forms of estrogens are water soluble and do not bind transport proteins and are therefore readily excreted -ia bile, feces and urine. Free estrogens are distributed to the bile, reabsorbed in the GI tract, and recirculated to the lier. About 20% of the estrogen is excreted in feces and the rest in urine (Piferrer, 2011). Estradiol is rapidly cleared from the blood whereas estrone is conerted in the lier to estrone sulfate which is excreted or hydrolyzed back to estrone. The serum half -life for estrone sulfate is about 12 hours and therefore its presence in water enironment is indicatie of continuous discharge of large quantities of untreated sewage into the lake. Howeer the common synthetic steroidal estrogens; ethynylestradiol and mestranol, are metabolized more slowly than estradiol because of the ethynyl group at C17 (Harding et al., 2013).
2.9 Detection of EDEs in water matrix
Estrogenic hormones in water matrix are usually determined by techniques such as gas chromatography, mass spectrometry, and high performance liquid chromatography. These methods are reliable but they hae seeral pitfalls, such as expensie instrumentation, complex deriatization, extensie clean up, and purification, and they also require a -ery high leel of technical expertise for operation. Howeer, the aailability of rapid, simple and cost effectie methods for quantitatie analysis of estrogenic hormones such as ELISA has downsized the use of these techniques. In addition, ELISA technique can be used to analyze a large number of samples simultaneously and machines that do the readings are relatiely cheap and also aailable in portable formats that can be used in field based studies (Swart & Pool, 2007).
In analysis of estrogens, ELISA kits operate on the basis of competition between an enzymelabelled conjugate and the free estrogen in the sample for a limited number of binding sites on the antibody coated plate. When properly -alidated for the species under study, ELISA technique is sensitie with the detection limit of 10 pg/mL and specific with minimum cross reactiity with other endocrine disrupters. Specifically, the estrogen monoclonal antibodies bind exclusiely with E1, E2 and E3 and do not cross react with other chemicals of similar structures. Proided the amount of labeled EDE is held constant, increasing the amount of unlabeled EDE reduces the amount of labeled EDE that is bound, thus resulting in a decrease in the intensity of color deelopment detected (Holland, 2003). Therefore, the extent of color deelopment is inersely proportional to the amount of EDE in the sample. The higher the concentration of the EDE in the sample, the less colour reaction produced and recorded using a standard microplate reader. The standard cure deeloped using known amounts of unlabeled estrogen is used to estimate the concentration of specific estrogen in the water samples. Therefore, the use of ELISA kits in the detection and quantification of estrogens in water in this study was premised on the principal adantages of ELISA techniques.
2.10 Biological and biochemical endpoints for ealuation of fish health and fecundity
The knowledge of reproductie biology of fish is a prerequisite of fish production. Thus, different biological and biochemical endpoints hae been exploited in the ealuation of reproductie potential and the oerall health of fish (Zin et al., 2011). The most biologically pertinent endpoints measured in an endocrine screening assay are those directly associated with reproduction: fecundity, fertilization success and embryo hatching. These endpoints are ideally quantified before and during the exposure period. Additional endpoints collected at termination of the exposure include Gonadosomatic Index (GSI), Hepatosomatic Index (HSI), Lengthweight Relationship (LWR), morphology, and biochemical endpoints such as -itellogenin (Vtg) and sex steroids (Folmar et al., 2000). In recent years, histological techniques hae gained -alue in the ealuation of oerall fish health in heaily polluted water enironments (Leino et al., 2005). The rationale and applications of the aboe endpoints are further elucidated in the sections that follow below.
2.10.1 Gonadosomatic index and hepatosomatic index
The condition and tissuesomatic indices, such as the gonadosomatic index (GSI) and hepatosomatic index (HSI), are a general measure of the oerall condition of the fish or growth status of a specific tissue. Tissuesomatic indices are commonly reported in fisheries studies because of the relatie ease of determination and the general belief that certain indices, such as the lier-somatic index, can be excellent predictors of aderse health in fish (Leino et al., 2005).
The GSI is also frequently reported as a general measure of gonad maturation and spawning readiness and is based on the broad assumption that proportionally larger gonads indicate greater deelopment. GSI -aries between males and females during reproduction cycle and it is influenced by the seasonal changes of the abiotic parameters such as temperature and photoperiod (Louiz et al., 2009). Thus, GSI is generally indicatie of reproductie success and general water quality. The GSI is potentially useful in reproductie screening, because reduction in relatie gonad mass can occur as a response to certain types of water pollutants (Leino et al., 2005). Gonadosomatic index measurements base on an assumption that linear relationships between gonad weight and body weight are constant throughout -arying stages of gonadal deelopment. The general procedure for determining the GSI is simple and inol-es euthanizing the fish, determining the total body mass, and then remoing and weighing the gonads. The index is then calculated as GSI = 100 × gonad weight (g) / body weight (g). In case of O. niloticus, GSI of 2.85 % for females and 0.41 % for males are considered as normal -alues (Akpu & Chindah, 2009). In addition, Khalafalla et al. (2010) reported that the normal GSI for O. niloticus ranges from 2.85 to 3.23 % for females and 0.98 to 1.36 % for males.
Howeer in crosssectional studies, this criterion is challenging to meet since it is quite rare to find female fish species that are synchronous or fractional spawners and the interindi-idual -ariation in oarian weight can be high during the spawning cycle. Another problem with the use of GSI is that normalizing gonad mass to body weight introduces bias into the analyses that could potentially lead to misinterpretation of the effect of toxicants on the size of the reproductie organs (Leino et al., 2005).
In a separate dimension, the lier is a metabolic organ and a site of most metabolic reactions in fish and therefore, HSI is a useful biomarker in the assessment of effects of hazardous enironmental stressors on the lier. Hepatosomatic index is defined as the percentage ratio of the lier weight to the body weight and proides an indication of the status of energy resere in a gien animal. It is calculated as HSI = 100 x lier weight (g) / body weight (g). In a polluted enironment, fish usually hae a smaller lier- with less energy resere (Akpu & Chindah, 2009). It has been reported that the normal HSI for O. niloticus ranges from 1.95 to 2.20% for both sexes (Khalafalla et al., 2010).
2.10.2 Histological techniques
During histological preparation of reproductie and other tissues of fish to assess toxicant effects of estrogens, paraffin embedding is recommended. Howeer, glycolmethacrylate embedding is a more specialized technique that results in the ability to more finely differentiate cellular and tissue structure (Holland, 2003). In addition, a standard chemical fixatie should be used for ealuations as it affects the degree of tissue shrinkage and alter staining properties of tissues. The recommended fixaties include but not limited to neutral buffered formalin, Bouin’s fixatie and buffered glutaraldehydeformaldehyde fixaties in regard to fixation of fish tissues. To minimize the autolysis and ensure satisfactory results, tissues should be placed in the fixatie immediately after excision from the specimens and adequate fixation time (at least 3 days) allowed prior to initiation of the histological matrix infiltration and embedding procedures (Jensen et al., 2001).
Although histology is a qualitatie to semiquantitati-e tool used to detect, describe alterations, and localize specific changes in -arious tissues, the method has been recognized as the most accurate for staging reproductie deelopment in fish (West, 1990). The primary difficulty in applying histological analysis is that interpretation of changes in tissues may -ary from one inestigator to another. Therefore, to maximize the repeatability of interpretation, indiidual histological slides should be blind coded to decrease bias. Besides, the method can be used to compare normal with abnormal tissue structures (Holland, 2003).
2.11 Histopathological effects of EDEs in fish
Fish are important -ehicles for the transfer of contaminants to human populations and may indicate the potential exposure to pollutants. Conersely, histopathological studies, in the laboratory and in field experiments, hae proed to be a sensitie tool to detect direct toxic effects of chemical compounds within target organs of fish (Abdel-Moneim et al., 2012). Estrogenic endocrine disrupters cause a wide range of aderse effects on the tissues/organs of fish. Notable potential effects of EDEs include decreased testicular growth and presence of primary egg cells in the testicular tissue (Prado et al., 2011). Howeer, histopathological alterations -ary according to the nature, amount, and route of entry of a gien e-DC and the tissue or organ affected. It has been reported that less than 1 ng/L of 17α-ethynylestradiol caused feminization of male fishes and 4 ng/L caused abnormal reproductie deelopment in male fathead minnows (Wright-Walter & Volz, 2009).
2.11.1 Gills
The gills of freshwater fish are the largest fraction of the total body surface area, which is in direct contact with the water. The complexity and constant contact with the surrounding water make the gill the first target to waterborne pollutants. Thus, changes in fish gills are among the most commonly recognized responses to enironmental stressors and are indicatie of physical and chemical stress (Abdel-Moneim et al., 2012). Therefore, the presence of edematous filaments together with intense lamellar -asodilatation and accumulation of fused pigments in -arious filamentous spaces are the common histopathological findings following prolonged exposure of gills with -arious water chemical pollutants including EDEs (Srijunngam & Wattanasirmkit, 2001).
2.11.2 Lier
The lier plays an important role in -ital functions of metabolism, and it is a major organ of accumulation, biotransformation, and excretion of contaminants of fish. Releant to this study, histopathological changes in liers of fish including O. niloticus exposed to a wide range of organic compounds and heay metals hae been reported (Abdel-Moneim et al., 2012). Besides, the normal lier cell structure exhibits a homogenous cytoplasm around the centrally located spherical nucleus. Howeer, marked cytoplasmic -acuolization, presence of hemosiderinladen macrophages as well as melanophages are the common findings associated with EDEs exposure. Other histopathological alterations include lier congestion resulting from the interruption of blood sinusoids that separate the hepatic cords, necrosis and presence of siderotic nodules. On the other hand, chronic exposure of fish to EDEs results into fatty lier degeneration associated with fibrosis caused by infiltration of collagen fibres in the hepatocytes (Srijunngam & Wattanasirmkit, 2001).
2.11.3 Kidneys
In fish, the spleen and kidneys hae important functions in implementing immune defense and the maintenance of a stable internal enironment. Histological changes proide a rapid method for detecting the effects of irritants in kidneys especially chronic ones (Wang et al., 2013). It has been reported that infiltration of the kidney tissue with numerous leukocytes, necrosis, and loss of structure are some of the symptoms associated with prolonged exposure of kidneys to waterborne chemical pollutants. Besides, the atrophy of kidney tubules as a result of hypersecretion of mucous cells is particularly associated with EDE exposure (Srijunngam & Wattanasirmkit, 2001). In addition, the presence of -arious siderotic nodules in the kidney is another common sign associated with EDE intoxication (Abdel-Hameid, 2007).
2.11.4 Muscle
Histopathological changes commonly associated with -arious waterborne chemical compounds in fish muscles include the disappearance of muscle striations, necrosis, and hyperplasia (Srijunngam & Wattanasirmkit, 2001). Muhamed (2008) reported that -acuolar degeneration in muscle bundles, splitting of muscle fibres and atrophy of muscle bundles were associated with chronic exposure of fish to -arious chemical water pollutants.
2.11.5 Testes and oaries
Many chemical compounds exert harmful effects on the reproduction of aquatic populations while interfering with the physiological processes. It has been reported that chronic exposure to pollutants induces a gonadal deterioration as determined by a reduction of GSI as well as morphological and histological changes of the gonads (Louiz et al., 2009). The effects of estrogens on male reproductie follicles include degeneration of spermatocytes, generalized organ atrophy, deelopment of oa-testis, and the proliferation of sertoli cells. Howeer, in females, atretic oocytes and abnormal formation of degeneratie oocytes are the main effects of EDEs (Leino et al., 2005).
CHAPTER THREE
3.0 MATERIALS AND METHODS
3.1 Research design
This was a crosssectional study that inol-ed both qualitatie and quantitatie data collection methods. Quantitatiely, the study inol-ed determination of estrone (E1), 17β-estradiol (E2), 17α-ethynylestradiol (EE2) and total estrogen (E1, E2, EE2) concentrations in water samples as well as physicochemical properties of water to assess quality of water in the sampling area. On the other hand, qualitatie methods inol-ed identification and description of pathological changes in tissues of O. niloticus from Napoleon Gulf of Lake Victoria, Uganda. Tissues from the lier, spleen, testis and oaries of O. niloticus were collected, processed and microscopically examined for any histopathological alterations . Sample collection and laboratory analysis were done between the months of May and June, 2013. Simultaneously, water samples were also collected from three sampling points along Lubigi channel in Kampala for comparatie estrogen analysis. A total of 33 water samples were collected from the study sites associated with -arious human actiities in the Napoleon Gulf of Lake Victoria. Concurrently, 72 tissue samples of fresh fish purchased from fishermen were analyzed for histopathological changes. To determine any crossreacti-e compound present in water samples, methanol solution (10 % -/-) was used as a blank. To crosscheck the extraction efficiency of the procedure, two liters of distilled water were mixed with known concentrations of total estrogens and equally analyzed for estrogens.
3.2 Sampling techniques
Purposie sampling was used to select the study sites. Initially, a surey was done to establish the potential sources of estrogenic endocrine disruptors in the study area. The sampling sites were selected basing on the leel of anthropogenic actiities. Using this criterion, eleen sampling sites and three sampling sites in Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala respectiely were selected for this study (Appendix VI). At eery study site, three sampling points were selected to take into account the source and effect of dilution on the concentration of the putatie estrogens in water. These were shoreline sampling point (0 m distance), intermediate sampling point (100 m distance from the shoreline), and a sampling point at a distance of 500 m from the shoreline.
Table 3: Sampling site descriptions in the study areas
Abbildung in dieser Leseprobe nicht enthalten
3.3 Location of the sampling sites
The Napoleon Gulf, located in the Jinja area of Lake Victoria comprises of Masese, Walukuba, and Kakira landing sites as the main fishing -illages on shores of Lake Victoria in Jinja Municipality. The study area is located on the eastern side of Jinja Municipality. The major islands in Napoleon Gulf include Rwebitooke, Kisima I and Kisima II. Masese, the major landing site has about 10,000 residents constituting 1420 households. Masese largely has unplanned human settlements, with poor social infrastructure. There are no sewer systems, and the large human population use pit latrines, bushes or the lake water as waste disposal points. The Napoleon Gulf was chosen for this study because it constitutes important fishery resources but also acts as waste discharge points for Jinja municipal wastewater (Report, 2010).
As a comparatie study site, the Lubigi channel located in Kawempe and Kampala central diisions in Kampala district was selected as another sampling site. This comparatie study area was chosen because it is a mirror reflection of what takes place in any other urban setting in Uganda such as Jinja Municipality. In addition, the channel has seeral discharge points for untreated human wastewater. It is surrounded by predominantly low income class of people who reside in semipermanent structures. Similar to Napoleon Gulf, most households in this area do not hae toilet facilities and discharge their waste directly into the channel (Matagi, 2002). It is therefore anticipated that this could be another potential source of EDEs like the Napoleon Gulf sites on Lake Victoria.
Abbildung in dieser Leseprobe nicht enthalten
Adapted from Linda et al. (2004)
Figure 2: Location of the Napoleon Gulf, Lake Victoria showing the study areas
3.4 Assessment of water quality at sampling points in Napoleon Gulf, Lake Victoria
The water quality at the study sites was assessed by determination of the physicochemical parameters such as pH, electrical conductiity, and temperature that were measured In situ using a field based three- in one CyberScan PC 300 meter. Measurement of dissoled oxygen (DO) was done using submersible DO meter. The meter was first calibrated using standard buffers according to the manufacturer’s instructions. Conersely, the CyberScan PC 300 meter was calibrated using pH buffer standards 4.0, 7.0, and 10.0. During the In situ measurements, the PC 300 meters were inserted into the water for at least 60 seconds before taking a reading to allow enough time for equilibration. For eery physicochemical parameter, three readings were taken at each of the sampling sites. The mean of the readings for each parameter was calculated and used as an indicator of water quality. Temperature was expressed in degrees Celsius (oC), DO in milligrams per liter (mg/L), and electrical conductiity in microsiemens per centimeter (mS/cm).
3.5 Collection of water and fish samples for analysis
Briefly, from eery study site, three water samples (2.5 L each) were collected from each of the three sampling points in a thoroughly cleaned amber bottle. Sampling bottles were appropriately labeled for identification. The water samples were immediately transferred into ice boxes and transported to the laboratory at the College of Veterinary Medicine, Animal Resources and Biosecurity, Makerere Uniersity. They were stored in fridge at 4oC until extraction of estrogens within a period of 48 hours after sample collection. Concurrently, wastewater samples were collected from three sampling points along Lubigi channel in Kampala for a comparatie study. One sample from a discharge point in Katanga, the second water sample from Kalerwe and the third water sample from Bwaise.
Simultaneously, fish samples of O. niloticus caught from or near the water sampling sites were procured from fishermen. This fish species was chosen because it has already been used in histopathological studies so background information is aailable. Each fish sample was weighed, examined for skin and gill lesions. Both the total length and standard length of fish were measured before dissection on a clean board using stainless knife. The lier, spleen, testes and oaries of the dissected fish were carefully remoed, weighed, and immediately transferred into sample bottles containing Bouin’s fixatie. The fish tissues were then transported to the Central Diagnostic Laboratory at Makerere Uniersity. After 24 hours of fixation, the tissues were remoed from the Bouin’s solution and washed seeral changes with 70% ethanol until clear. The washed fish tissues were then kept in 10% neutral buffered formalin solution until processed for histopathological examination.
3.6 Isolation of estrogens from water samples
To isolate estrogens from water, the samples were subjected to C18 Solid Phase Extraction (SPE) columns (BakerbondTM) according to the protocol described by Swart and Pool (2007) with minor modifications customized to our laboratory settings. Briefly, before extraction on C18 SPE columns (BakerbondTM), two liters of each water sample were filtered through cotton wool and thereafter the filtrate was passed through a glass microfiber filter (47mm in diameter- Cat No. GF3047). Thereafter, glass fiber filtrates were subjected to C18 SPE adsorption chromatography. Initially, C18 SPE columns (BakerbondTM) were prewashed with four milliliters of the solent mixture (40% hexane, 45% methanol, and 15% 2-propanol), followed by another wash with four milliliters of absolute ethanol. The columns were then washed with six milliliters of double distilled water after which the filtered water samples were applied onto the columns. The columns were then air dried with a manual -acuum pump until dry. The bound hydrophobic substances were eluted with fie milliliters of the solent mixture. The eluate was then reconstituted in two milliliters of absolute ethanol and used in ELISA detection of estrogens. The eluates were stored at -200C until analysis. To ealuate the extraction efficiency of the extraction procedure, two liters of distilled water were mixed with known concentrations of total estrogens. During filtration and extraction, water samples were regularly shaken to ensure use of a homogenous solution.
Abbildung in dieser Leseprobe nicht enthalten
Figure 3: Flow chart of Solid Phase Extraction (SPE) of estrogens from water samples.
3.7 ELISA determination of estrogen concentrations in water samples
The concentrations of total estrogens (E1, E2 & EE2), 17β-estradiol (E2), estrone (E1), and 17α-ethynylestradiol (EE2) were determined using specific ELISA kits (Ecoloqiena(R)) from Japan Eniro Chemicals, Ltd. Before analysis, the extracted estrogens in form of concentrated eluates were diluted 1/10 using distilled water and the antigenenzyme solution was reconstituted according to the manufacturer’s recommendations. The reconstituted antigenenzyme solution was stored at 4oC and used in two days. The diluted eluates were then assayed directly with specific ELISA kits (Ecoloqiena(R)) in duplicates.
3.7.1 Determination of total estrogens
Briefly, 100µl of either the extracted water sample or the standard total estrogens and 100µl of the reconstituted antigenenzyme solution were mixed in each well of the uncoated microplate using a microtiter pipette. One hundred microliter aliquots of the aboe mixture were dispensed into each of the precoated wells of the microplate. The contents of the wells were gently mixed by pipetting in and out. The ELISA microplate was then incubated at room temperature (250C) for 60 minutes. After incubation, the wells were washed three times using 300µl of a sixfold concentrated wash solution. After proper washing, 100µl of the TMB were added into each well and incubated at room temperature (250C) for 30 minutes. After incubation for 30 minutes, the reaction was stopped by addition of 100µl of 4M H2SO4 into each of the wells. The absorbance of the colour deeloped was measured within 15 minutes at 450 nm using a microplate reader (Gene 5). A standard cure was drawn using ELISA absorbance readings of the standards (0, 0.05, 0.15, 0.5 and 3.0µg/L prepared in 10% methanol). The concentrations of total estrogens in each sample were estimated from the standard cure.
3.7.2 Determination of estrone
Briefly, 100µl of either the extracted water sample or the standard estrone and 100µl of estronehorseradish peroxidase (HRP) conjugate were mixed in each well of the uncoated microplate using a microtiter pipette. One hundred microliter aliquots of the aboe mixture were dispensed into each of the rabbit antiestrone precoated wells of the microplate. The contents of the wells were gently mixed by pipetting in and out. The ELISA microplate was then incubated at room temperature (250C) for 60 minutes. After incubation, the wells were washed three times using 300µl of a sixfold concentrated wash solution. After thorough washing, 100µl of the TMB substrate were added into each well and incubated at room temperature (250C) for 30 minutes. After incubation for 30 minutes, the reaction was stopped by addition of 100µl of 4M H2SO4 into each of the wells. The absorbance of the colour deeloped was measured within 15 minutes at 450nm using a microplate spectrophotometer (Gene 5). A standard cure was drawn using ELISA absorbance readings of the standards (0, 0.05, 0.3, 0.8 and 5.0µg/L prepared in 10% methanol). The concentrations of estrone in each sample were estimated from the standard cure.
3.7.3 Determination of 17β-estradiol
Briefly, 100µl of either the extracted water sample or the standard estradiol and 100µl of the 17β-estradiol –HRP conjugate were mixed in each well of the uncoated microplate using a microtiter pipette. One hundred microliter aliquots of the aboe mixture were dispensed into each of the precoated wells of the microplate. The contents of the wells were gently mixed by pipetting in and out. The ELISA microplate was then incubated at room temperature (250C) for 60 minutes. After incubation, the wells were washed three times using 300µl of a sixfold concentrated wash solution. After thorough washing, 100µl of the TMB substrate were dispensed into each well and incubated at room temperature (250C) for 30 minutes. After incubation (30 minutes), the reaction was terminated by adding 100µl of 4M H2SO4 into each of the wells. The absorbance of the color deeloped was measured within 15 minutes at 450 nm using a microplate spectrophotometer (Gene 5). A standard cure was drawn using ELISA absorbance readings of the standards (0, 0.05, 0.15, 0.4 and 1.0µg/L prepared in 10% methanol). The concentrations of 17β-estradiol in each sample were estimated from the standard cure.
3.7.4 Determination of 17α-ethynylestradiol
Briefly, 100µl of either the water extract or the standard 17α-ethynylestradiol and 100µl of the 17α-ethynylestradiol –HRP conjugate were mixed in each well of the uncoated microplate using a microtiter pipette. One hundred microliter aliquots of the aboe mixture were dispensed into each of the precoated well of the microplate. The contents of the wells were gently mixed by pipetting in and out. The ELISA microplate was then incubated at room temperature (250C) for 60 minutes. After incubation, the wells were washed three times using 300µl of a sixfold concentrated wash solution. After thorough washing, 100µl of the TMB substrate were added into each well and incubated at room temperature (250C) for 30 minutes. After incubation, the reaction was stopped by addition of 100µl of 4M H2SO4 into each of the wells. The absorbance of the colour deeloped was measured within 15 minutes at 450nm using a microplate spectrophotometer (Gene 5). A standard cure was then drawn using ELISA absorbance readings of the standards (0, 0.05, 0.15, 0.5 and 3.0 µg/L prepared in 10% methanol). The concentrations of 17α-ethynylestradiol in each sample were estimated from the standard cure.
3.8 Calculation of gonadosomatic and hepatosomatic indices of O. niloticus
Lie fresh fish purchased from fishermen were weighed, euthanized and carefully dissected to remoe all the internal organs. The weights of the kidney, lier and the gonads were also measured in grams. The gonadal somatic index (GSI) and hepatosomatic index (HSI) were calculated from the formula GSI= 100 x gonad weight (g)/body weight (g) and HSI= 100 x lier weight (g)/ body weight (g) respectiely.
3.9 Preparation of fish tissue samples for histopathological analysis
The preparation of the O. niloticus tissue samples for histopathological analysis was done as per the recommended procedures (Dumitrescu et al., 2010). Fish tissue samples were processed after 3 days of fixation first in Bouin’s solution for 24 hours and later in 10% neutral buffered formalin for the remaining two days.
3.9.1 Sectioning of fixed samples
Briefly, Bouin’s and 10% neutral buffered formalin fixed organs were trimmed using a scalpel blade to approximately 5 mm thick in transerse sections and then placed in labeled tissue cassettes for processing and embedment.
3.9.2 Tissue processing and staining
Briefly, the cassettes containing sectioned tissues (5 mm thick) were placed in an automated tissue processor and passed through increasing concentrations of alcohol to dehydrate that is 80% alcohol for four hours and then 94% alcohol two changes for two hours. The tissues were then treated through three changes of isopropanol for 90 minutes each and the alcohol was cleared using amyl acetate for one hour. Subsequently, the tissues were treated in two changes of xylene for two hours and two and half hours respectiely. The samples were then impregnated in two changes of molten wax at 600C for three hours followed by embedding in molten wax in tissue cassettes. Fie-micrometer thick sections were cut using a rotary microtome and mounted onto labeled glass slides, and then allowed to dry oernight. The sections were then dewaxed in two changes using xylene each for fie minutes, followed by rehydration in alcohol starting with two changes of absolute ethanol to 95% ethanol for fie minutes each, and then water for fie minutes. The sections were then stained with Mayer’s hematoxylin for ten minutes then immersed in water for 15 minutes, during which, the nucleus would stain blueblack. Subsequently, the tissues were counterstained with 1% eosin for four minutes. In this stage, the rest of the cellular organelles would stain pink to red. The excess stain was washed off and tissues dehydrated using increasing concentrations of alcohol (two changes), cleared using two changes of xylene for two minutes each and mounted onto labeled glass slides with Dibutyl Phthalate Xylene (DPX) mountant.
3.9.3 Microscopic obseration of heamatoxylin and eosin stained tissues
Tissues were examined under light microscope at x 4, 10, 40, and 100 by two independent personnel with good knowledge of the histological changes in fish. The lesions obsered were recorded according to the tissue affected. Photomicrographs were also taken to document the obsered tissue pathologies.
3.10 Data recording
All samples and data collected were recorded in specially designed data collection sheets. Numerical data obtained from this study were entered in Microsoft Excel Sheet, 2007. Meanwhile, qualitatie data were described and recorded in a table form. In addition, photomicrographs of histopathological examinations were taken and presented as figures.
3.11 Data Analysis
Data collected were analyzed using Microsoft Excel 2007 and Graph Pad Prism 6. Graphical presentations of the means and standard error of the mean (SEM) was done using Microsoft Excel 2007 and Graph Pad Prism 6. Differences between the means of estrogens were analyzed by oneway analysis of -ariance (ANOVA). A statistical leel of significance of α= 0.05 was chosen and Tukey’s Honestly Significant Difference (HSD) test was used for multiple comparisons of means.
3.12 Ethical issues
Fish samples procured for dissection were euthanized before organs were collected. This was done by seering the spinal cord behind the head.
CHAPTER FOUR
4.0 RESULTS
4.1 Water quality Parameters
To assess the quality of water in the study area, selected physicochemical parameters of water were determined (Table 4). These include; pH, temperature, dissoled oxygen, and electrical conductiity. In general, the mean pH -alues of water increased along the sampling gradient. The mean pH -alue was lowest at 0 m (shoreline), intermediate at 100 m distance from the shoreline and was highest at 500 m distance from the shoreline (Figure 4A). Howeer, a statistical significant difference in the mean pH -alues was only recorded between the two categories of water that is at a distance of 0 m and 500 m distance from the shoreline (p < 0.05). Correspondingly, temperature readings decreased along the sampling gradient. The temperatures were highest at 0 m distance, intermediate at 100 m distance off shoreline and a similar mean temperature -alue was determined at 500 m distance off shoreline (Figure 4B). Multiple comparisons of mean temperature -alues among the two categories of water (water samples from 100 m distance and 500 m distance) showed no significant statistical difference (p > 0.05). Howeer, there was a significant difference between the water samples from 0 m and 500 m as well as 100 m distances from the shoreline (p < 0.05). On the other hand, the mean concentration of dissoled oxygen increased along the sampling gradient being lowest at 0 m distance, intermediate at 100 m distance off shoreline, and highest at 500 m distance off shoreline ( Figure 4C). The multiple comparisons of the mean dissoled oxygen -alues showed a significant difference between two categories of water that is at a distance of 0 m and 500 m distance off shoreline respectiely (p < 0.05). Additionally, the concentration of dissoled ions/EC generally decreased along the sampling gradient. The mean electrical conductiity -alues determined were highest at a distance of 0 m and the lowest at 500 m distance off shoreline (Figure 4D). Howeer, there was no statistical difference among mean electrical conductiity -alues among the three categories of water (p > 0.05).
Table 4: The mean -alues of physicalchemical parameters of water at different sampling points in Napoleon Gulf, Lake Victoria
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Data are presented as mean + standard error of the mean Abbildung in dieser Leseprobe nicht enthalten Figure 4: Water quality parameters at different sampling points in Napoleon Gulf of Lake Victoria. A; pH, B; temperature, C; DO, and D; electrical conductiity at different sampling sites.
4.2 Concentration of total estrogens in water
Analysis of water samples from different sampling points showed -ariations in the concentration of total estrogens. The detection leels of the assay for total estrogens in the study ranged from 200 to 1050 ng/L (Table 5). In the Napoleon Gulf, the highest concentration of total estrogens was detected in water samples from Masese landing site at 0 m distance (800 ng/L) whereas the lowest concentration was detected in water samples from Wanyange drainage at 500 m distance from the shoreline (200 ng/L). With exception of water samples from Kakira and Kirinya –leather industry, the concentration of total estrogens decreased along the sampling gradient in the Napoleon Gulf. The highest concentrations of total estrogens were detected in water samples from the shoreline, intermediate concentrations at 100 m distance from the shoreline and lowest concentrations in water samples from 500 m distance from the shoreline (Figure 5). Multiple comparisons of total estrogens concentrations indicate no significant difference (p > 0.05) among the three categories of water samples. In comparison, water samples from Lubigi channel had higher concentrations of total estrogens than water samples from the Napoleon Gulf of Lake Victoria except for water samples from Katanga. From Kampala, the highest concentration of total estrogens was detected in water samples from Bwaise (1050 ng/L), followed by Kalerwe (1000 ng/L) and then Katanga (463 ng/L) as shown in Table 5.
Table 5: Concentrations of total estrogens (ng/L) in water samples from different sampling points in Napoleon Gulf of Lake Victoria and Lubigi Channel in Kampala, Uganda
Abbildung in dieser Leseprobe nicht enthalten
a: concentration of total estrogens in a water sample from Katanga, the origin of the channel, b: concentration of total estrogens in a water sample from Kalerwe, a sampling point at 100 m from Katanga, and c: concentration of total estrogens in a water sample from Bwaise sampling point that is about 500 m from Katanga and is downstream of Katanga and Kalerwe sampling points. Generally, the concentration of total estrogens in water samples from Lubigi Channel was higher than in water samples from Napoleon Gulf.
Abbildung in dieser Leseprobe nicht enthalten Figure 5: Concentrations of total estrogens (ng/L) in water samples from different sampling points in Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda.
4.3 Concentration of estrone in water
Low leels of estrone were detected in water samples from Napoleon Gulf of Lake Victoria (Table 6). No estrone was detected in water from some sampling sites especially those at 500 m distance from the shoreline (Table 6). The detection leels of estrone in water samples ranged from 9.8 to 49 ng/L. There was a significant difference in estrone concentrations between water samples from shoreline (0 m distance) and 500 m distance off shoreline sampling points (p < 0.05). Besides, the highest concentration of estrone was recorded in water samples from a sampling site presumed to hae no human actiities (49 ng/L) while the lowest concentration was detected in water samples from Bidco and soap industry discharge point at the shoreline (9.8 ng/L). No estrone was detected in water samples from the different sampling points at Wanyange landing site (Table 6). With exception of Bidco, the concentration of estrone in water samples from other sampling sites decreased along the sampling gradient in the Napoleon Gulf (Figure 6). On the other hand, water samples from Lubigi channel had considerably higher concentrations of estrone than water samples from Napoleon Gulf of Lake Victoria. In addition, the highest concentration of estrone was detected in water samples from Kalerwe (750 ng/L), followed by Bwaise (730 ng/L), and then Katanga (450 ng/L) as shown in table 6.
Table 6: Concentrations of estrone (ng/L) in water samples from different sampling points in Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
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ND = Not Detected. a: concentration of estrone in a water sample from Katanga, a point where the channel starts from, b: concentration of estrone in a water sample from Kalerwe, a sampling point at 100 m distance from Katanga, and c: concentration of estrone in a water sample from Bwaise sampling point that is about 500 m distance from Katanga and is downstream of Katanga and Kalerwe.
Abbildung in dieser Leseprobe nicht enthalten
Figure 6: Concentrations of estrone (ng/L) in water samples from Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda. The concentration of estrone was highest in water samples from the shoreline (0 m distance) and lowest in water samples from 500 m distance from the shoreline.
4.4 Concentration of 17β-estradiol in water
In comparison, the leels of 17β-estradiol detected in water samples from Napoleon Gulf, Lake Victoria were higher than estrone and 17α-ethynylestradiol (Figure 8). The 17β-estradiol detection leels in water samples ranged from 145 to 305 ng/L. The 17β-estradiol concentrations in water samples decreased along the sampling gradient (Figure 6). The 17β-estradiol concentrations detected in water samples were highest at shoreline sampling point (0 m distance), intermediate at 100 m distance from the shoreline, and lowest at 500 m distance from the shoreline (Figure 6). Howeer, there was a significant difference in the 17β-estradiol concentrations detected in the three categories of water samples (p < 0.05). The highest 17β-estradiol concentration (305 ng/L) was detected in a water sample from the shoreline at Masese Landing site whereas the lowest concentration of 17β-estradiol (145 ng/L) was detected in a water sample from Wanyange drainage at 500 m distance from the shoreline (Table 7).
On the other hand, water samples from Lubigi channel had higher concentrations of 17β-estradiol than water samples from Napoleon Gulf except for a water sample from Katanga. The highest concentration of 17β-estradiol was detected in a water sample from Bwaise (329 ng/L), followed by Kalerwe (318 ng/L), and then Katanga (275 ng/L) as shown in table 7.
Table 7: Concentrations of 17β-estradiol (ng/L) in water samples from different sampling points in Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda
Abbildung in dieser Leseprobe nicht enthalten
a: concentration of 17β-estradiol in a water sample from Katanga, a point where the channel starts from, b: concentration of 17β-estradiol in a water sample from Kalerwe, a sampling point at 100 m distance from Katanga, and c: concentration of 17β-estradiol in a water sample from Bwaise sampling point that is about 500 m distance from Katanga and is downstream of Katanga and Kalerwe sampling points. From the Napoleon Gulf, Masese landing site had the highest concentration of 17β-estradiol (305 ng/L)** at 0 m distance while Wanyange drainage with mixed industrial wastewater had the lowest concentration of 17β-estradiol (145 ng/L)*** at 500 m distance from the shoreline.
Abbildung in dieser Leseprobe nicht enthalten
Figure 7: Concentrations of 17β-estradiol (ng/L) in water samples from Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda.
4.5 Concentration of 17α-ethynylestradiol in water
Compared to estrone concentrations determined in this study, slightly higher concentrations of 17α-ethynylestradiol were detected in water samples from the Napoleon Gulf of Lake Victoria. The detection leels for 17α-ethynylestradiol in water samples ranged from 45 to 360 ng/L (Table 8). The results in table 8 further indicate that water samples from Masese landing site had the highest concentration of 17α-ethynylestradiol (360 ng/L) at 100 m distance from the shoreline, samples from the National Water and Sewerage Cooperation in Jinja had the lowest concentration of 17α-ethynylestradiol (45 ng/L) at 500 m distance from the shoreline. In general, the mean 17α-ethynylestradiol concentrations in water samples decreased along the sampling gradient (Figure 7). There was a significant difference among the mean 17α-ethynylestradiol concentrations detected in water samples from the shoreline and 500 m distance from the shoreline sampling sites (p < 0.05) as indicated in figure 7. Conersely, water samples from Lubigi channel had lower concentrations of 17α-ethynylestradiol than water samples from Napoleon Gulf. The highest concentration of 17α-ethynylestradiol was detected in water samples from Bwaise (105 ng/L), followed by Katanga (82 ng/L), and then Kalerwe (47 ng/L) as shown in table 8.
Table 8: Concentrations of 17α-ethynylestradiol (ng/L) in water samples from different sampling points in Napoleon Gulf of Lake Victoria and Lubigi channel in Kampala, Uganda
Abbildung in dieser Leseprobe nicht enthalten
a: concentration of 17α-ethynylestradiol in a water sample from Katanga, a point where the channel starts from, b: concentration of 17α-ethynylestradiol in a water sample from Kalerwe, a sampling point at 100 m from Katanga, and c: concentration of 17α-ethynylestradiol in a water sample from Bwaise sampling point that is about 500 m from Katanga and is downstream of Katanga and Kalerwe. From the Napoleon Gulf, Masese landing site had the highest concentration of 17α-ethynylestradiol (360 ng/L)** at 0 m distance from the shore while municipal waste water from NSWC in Jinja had the lowest concentration of 17α-ethynylestradiol (45 ng/L)*** at 500 m distance from the lake shore.
Abbildung in dieser Leseprobe nicht enthalten Figure 8: Concentrations of ethynylestradiol (ng/L) in water samples from Napoleon Gulf, Lake Victoria and Lubigi channel in Kampala, Uganda. *p < 0.05 for 0 m Vs. 500 m except for Lubigi channel. Howeer, there was no significant difference in 17α-ethynylestradiol concentrations in water samples from 100 m distance from the shoreline and 500 m distance from the shoreline sampling points (p > 0.05).
4.6 Variations in estrogen concentrations in water samples from Napoleon Gulf
The concentration of total estrogens in water samples from Napoleon Gulf was higher than estrone (E1), 17α-ethynylestradiol (EE2) and 17β-estradiol (E2) concentrations. In a decreasing order, the concentration of total estrogens was followed by 17β-estradiol, then 17α-ethynylestradiol and finally estrone. In general, the concentrations of estrogens (total estrogens, 17β-estradiol, 17α-ethynylestradiol and estrone) decreased along the sampling gradient, being highest in water samples from the shoreline sampling point (0 m distance), intermediate in water samples from 100 m distance from the shoreline and lowest in water samples from 500 m distance from the shoreline (Figure 9).
Abbildung in dieser Leseprobe nicht enthaltenFigure 9: The -ariations in the concentration of total estrogens, 17β-estradiol and 17α-ethynylestradiol, estrone (μg/L) in water samples from Napoleon Gulf of Lake Victoria, Uganda
4.7 Endpoints in fish health and production
To ealuate the health and reproductie performance of fish caught from the study area, morphological, gonadosomatic index, hepatosomatic index and histopathological endpoints were assessed.
4.7.1 The gonadosomatic index (GSI)
The GSI of fish was used to assess the reproductie performance of both male and female fishes. Results show that female fish caught from different areas had a higher mean GSI (2.27 + 0.09 %) than their male counterparts (2.18 + 0.13 %) (Table 9). Among the female fish, those caught from Wanyange fishing area had the highest GSI (2.63 + 0.33 %), followed by Wairaka (2.4 + 0.22 %), Bidco (2.22 + 0.09 %), Masese (2.2 + 0.31 %), Kirinya (2.16 + 0.04 %), and Kakira (2.02 + 0.44 %). Conersely, male fish from Kakira had the least mean GSI (1.79 + 2.7 %), followed by Wairaka (1.94 + 0.6 %), Bidco (2.07 + 0.06 %), Kirinya (2.26 + 0.04 %), Wanyange (2.41 + 0.2 %), and Masese fishing area had male fish with the highest GSI (2.62 + 0.17 %) (Figure 10).
Table 9: Mean GSI (%) and HSI (%) of female and male fish from Napoleon Gulf of Lake Victoria, Uganda
Abbildung in dieser Leseprobe nicht enthalten
Key: W=Wairaka, M=Masese, Ka=Kakira, Wan=Wanyange, B=Bidco, and Ki=Kirinya sampling sites
Abbildung in dieser Leseprobe nicht enthalten
Figure 10: Mean GSI (%) of fish caught from Napoleon Gulf of Lake Victoria, Uganda. The reference ranges were 2.85-3.23% for females and 0.98- 1.36 % for males. All male fish caught from different areas of Napoleon Gulf had higher GSI compared to females caught from the same area.
4.7.3 The hepatosomatic index (HSI)
The HSI was used to assess the probable effects of chemical pollutants in water on the lier of fish. In general, there was a significance difference (p < 0.05) of HSI among fish from different sampling points but the female fish had higher mean HSI of 2 + 0.09 % compared to male fish (2.18 + 0.13 %) (Table 9). The results show that fish caught from Wanyange fishing area had the highest mean HSI of 3.14 % and 3.38 % for male fish and female fish respectiely. Male fish caught from Bidco and Wairaka fishing areas generally had lower mean HSI than their female counterparts caught from other sites (Figure 11).
Abbildung in dieser Leseprobe nicht enthalten
Figure 11: Mean HSI (%) of fish caught at different sites in the Napoleon Gulf of Lake Victoria, Uganda. The reference range was 1.95-2.2% for males and females. Both female and male fish caught in Napoleon Gulf of Lake Victoria hae high HSI, except those from Masese, Kirinya and Kakira areas.
4.8 Histopathological analysis
Histopathological changes in most of the fish samples were quite similar (Appendix VII). In the spleen tissue, seere congestion, focal areas of necrosis, mild tissue degeneration, presence of hemosiderin laden macrophages, siderotic nodules as well as apoptosis were the common histopathological changes obsered (Figure 12). In the lier tissue, focal areas of necrosis, congestion, and apoptosis were -isible (Figure 13). In the testicular tissue, extensie areas of necrosis, cell desquamation and degeneratie changes were obsered (Figure 14). In addition, the presence of atretic follicle was the major histopathological alteration obsered in oaries (Figures 15 & 16). The atretic oocytes were characterized by loss of oocytes shape, and a disorganized and fragmented yolk. The breakdown of granulose layers that maintain the structural integrity of the oocytes leads to structural impairment and functional loss.
Abbildung in dieser Leseprobe nicht enthalten
Abbildung in dieser Leseprobe nicht enthalten
Figure 12: Spleen tissue of O. niloticus. Siderotic nodules characterized by accumulation of hemosiderin laden macrophages (siderophores) were seen in the spleen. S; siderotic nodules and M; melanophages (x 100, Haematoxylin and Eosin).
Abbildung in dieser Leseprobe nicht enthalten
Abbildung in dieser Leseprobe nicht enthalten
Figure 13: Lier tissue of O. niloticus: LI; leukocyte infiltration and M; melanophage (x 100, Haematoxylin and Eosin).
Abbildung in dieser Leseprobe nicht enthalten
Abbildung in dieser Leseprobe nicht enthalten
Figure 14: Testicular tissue of O. niloticus. N; extensie areas of necrosis in the seminiferous tubules with cellular desquamation, and F; fibrosis in the walls of seminiferous tubules (x 100, Haematoxylin and Eosin).
Abbildung in dieser Leseprobe nicht enthalten
Abbildung in dieser Leseprobe nicht enthalten
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Figure 15: O-arian tissue of O. niloticus. Note the different deelopmental stages of follicles; Cn- chromatin nuclear oocytes; P –perinucleolar oocytes; Cacortical aleolar oocytes; V--itellogenic stage; R-ripe oocytes; Af- atretic follicles (x 100, Haematoxylin and Eosin).
Abbildung in dieser Leseprobe nicht enthalten
Figure 16: O-arian tissue of O. niloticus. Note the peculiarities in the different deelopmental stages; Cn- chromatin nuclear oocytes; P –perinucleolar oocytes; Cacortical aleolar oocytes; V--itellogenic stage; R-ripe oocytes; Afatretic follicles. Note the peculiar differences in the sizes of the follicles compared to Figure 15 (x100, Haematoxylin and Eosin).
CHAPTER FIVE
5.0 DISCUSSION OF RESULTS
5.1 Physicochemical parameters of water quality
The results of this study show that the mean pH -alues were low for samples collected from the shoreline and high in samples from off shoreline sampling points. Howeer, these -alues are generally higher compared to reported pH -alues of most water bodies. Kishe (2004) reported that although the pH of water bodies is influenced by the leel of human actiities, the pH of most water bodies ranges from 6.5 to 8. During sampling, discharge of untreated effluents in the study area was noted as well as the presence of algal blooms mainly at the shoreline sampling points. It was further obsered that the transparency of the water of shoreline and off shoreline sampling sites was poor indicating a high degree of pollution. The high pH obtained in this study could be attributed to presence of organic materials such as fertilizers in water. According to Birungi et al. (2007), humans are contributors of eleated water pH primarily in form of plant nutrient runoffs such as fertilizers that promote algal growth and also raise pH. Howeer, although studies hae shown that the optimum pH for most fish species is 7.4, O. niloticus suri-es in a wider pH range of 3.5 to 12 (Birungi et al., 2007). This obseration suggests that there was no significant effect of pH on health and production performance of O. niloticus in the area of study. Thus, the obsered reductions in the number of fish catches in the study area could be due to other factors than pH. Inappropriate fishing methods and practices such as capture of immature fish as well as oerfishing could explain the obsered reductions in fish catches. This conclusion is supported by Naigaga et al. (2011) who reported that significant declines in fish stocks and fish species diersity in Lake Victoria are due to oerexploitation, use of destructie fishing gear or methods, and ineffectie or absent management of human wastes. Conersely, the high pH -alues recorded at different sampling points in Napoleon Gulf might hae had a positie influence on the concentrations of estrogens detected in the different water samples. This obseration is supported by Shareef et al. (2006) who reported that pH influences the solubility of estrogens in water since the higher the pH the more ionized the estrogens and thus the higher the solubility and concentration in water.
According to Kishe (2004), the temperature of water influences the amount of dissoled oxygen and the water temperature depends on the season and time of sampling. The high temperature range recorded in this study could be attributed to the time of sampling. In the present work, collection of water samples was done during the dry season and between 9:00 AM and 2:00 PM and this might hae had a positie influence on the high temperature readings of water determined at different sampling points. Howeer, it could be concluded that water temperature in Napoleon Gulf of Lake Victoria probably had no impact on the reproductie performance of O. niloticus because the optimal temperature for normal growth of most tilapia species including O. niloticus ranges from 25 to 28oC (Hassan et al., 2013).
The dissoled oxygen (DO) concentration recorded in water samples from the shoreline was lower than the concentration in offshore water samples. The dissoled oxygen leels recorded in this study were higher (detection range of 3.11 to 15.7 mg/L) than those of other studies (Kishe, 2004; Sánchez et al., 2007). This was probably due to low leel of organic decomposition occurring in the deeper waters of the lake. In contradiction, the obsered algal blooms mainly at the shorelines during sampling indicate high leels of eutrophication. This obseration is supported by Kishe (2004) who obsered that the presence of high organic compounds in water such as fertilizers influences the degree of eutrophication and in turn the amount of DO in water; the higher the degree of eutrophication the lower the amount of DO in water. Besides, the DO in this study may not hae had significant effects on the reproduction of fish since the recommended DO for good growth is aboe 2 mg/L (Cleoni dos Santos et al., 2012).
The electrical conductiity of water depends on the leel of inorganic ions present, thus, the higher the concentration of organic compounds relatie to inorganic ions in water, the lower the conductiity (Hassan et al., 2013). The results of electrical conductiity (EC) in this study indicate small -ariations in water samples from the shoreline and offshore sampling points. The relatiely high EC -alues of water at the shoreline could be attributed to the discharge of effluents containing ions while the lower EC readings of offshore water could be a result of dilution by water currents. The comparatiely higher EC readings obtained in this study could be attributed to the discharge of mainly industrial wastes containing high leels of inorganic ions into the lake. According to Naigaga et al. (2011), the Napoleon Gulf of Lake Victoria receies discharges from domestic, surface runoffs, and industries from Jinja Municipality many of which contain high amounts of inorganic ions.
5.2 Total estrogens in water samples
Analysis of total estrogens in water samples from Napoleon Gulf, Lake Victoria showed higher leels of these hormones with a detection range from 200 to 800 ng/L compared to findings of other studies. This contradicts preious findings of low leels of total estrogens obsered in water samples from Mwanza Gulf, Tanzania (Mdegela et al., 2013) in which the detection leels of total estrogens ranged from 0.01 to 0.2 ng/L. This suggests that the Ugandan side of Lake Victoria is highly polluted with total estrogens than the Tanzanian side. According to Naigaga et al. (2011), the Napoleon Gulf of Lake Victoria and its wetlands receie greater quantities of untreated sewage and industrial wastes which are potential anthropogenic sources of estrogens. Additionally, the management of municipal wastes in Uganda is a presere of National Water and Sewerage Cooperation but the current sewer system in Kampala, the capital city of Uganda seres only a small fraction of the city population and only 10% of all sewage generated gets treated (Naigaga et al., 2011). Therefore, the detected high leels of total estrogens in water of Napoleon Gulf of Lake Victoria could be due to discharge of greater quantities of untreated Jinja municipal waste into the lake. During sampling, it was further obsered that humans residing in the study area use the lake for bathing and watering their animals and some toilets and urinals were situated at the shoreline and yet all these anthropogenic actiities are potential sources of EDEs. Furthermore, although waste water treatment efforts are in place in Uganda by the National sewerage Cooperation, this does not target the remoal of estrogens from influents. Thus, the EDEs may find their way into these water bodies endangering life of fish and humans.
Conersely, the concentration of total estrogens was higher in water samples from the shoreline (0 m) compared to samples from offshore points (100 m & 500 m). This finding was expected since the off shoreline water samples are likely to contain lower concentrations of enironmental pollutants compared to water samples from the shoreline probably a result of mixing and dilution by water currents. In addition, the high concentration of total estrogens in samples from Masese landing site could be attributed to the large number of animals in the area that urinate in water as well as human actiities such as bathing in the lake. Furthermore, due to the steep terrain and huge human population in the neighboring town, discharge of large amounts of untreated waste water into the lake and surface runoffs are more likely to occur.
In comparison, the presence of higher concentrations of total estrogens in water samples from Lubigi channel in Kampala compared to those from the Napoleon Gulf of Lake Victoria could be explained partly by the large human population who directly discharge untreated waste into the channel. It is known that Lubigi channel directly drains human and animal wastes (fecal and urine) from the densely populated suburbs of Kampala such as Kawempe, Bwaise, Katanga, Kikoni and Kalerwe. Most of the houses in these areas are located in low lying areas, and because most of the families in these areas lacked toilet facilities, they directly release wastes into the aailable drainage channels leading into the Lubigi channel. Additionally, located upstream of Lubigi channel, is the National Referral Hospital, Mulago from where some hospital wastes and untreated hospital sewage also flow as was eident during sample collection. These obserations are supported by Matagi (2002) who reported that most of the pollutants from anthropogenic actiities in Kampala endup in Lake Victoria through -arious channels draining the city. It is therefore anticipated that large amounts of untreated human wastes containing EDEs may find their way into Lake Victoria and other water bodies thus endangering fish life and that of other aquatic wildlife.
5.3 Estrone in water samples
Analysis of estrone in water samples indicated lower concentrations than 17β-estradiol and 17α-ethynylestradiol. The detection leel of estrone in samples ranged from 9.8 to 49 ng/L. This finding was expected since large amounts of estrone conjugates are hydrolyzed back to estrone in the lier and excreted in urine and feces in small quantities compared to 17β-estradiol (Minneman et al., 2005). Surprisingly, it was found that a water sample from the shoreline (0 m) at Kirinya sampling point had the highest leel of estrone (49 ng/L). This was not anticipated since this study area was presumed to be the least polluted in the whole of the Napoleon Gulf of Lake Victoria due to absence of eident anthropogenic actiities apart from occasional fishing. This finding would seem to imply that the whole of Napoleon Gulf is heaily polluted with untreated sewage. This obseration is supported by Swart and Pool (2007) who reported that 70 % of estrone, a natural hormone is excreted by both humans and animals through feces and urine and is present in water in large amounts through discharge of untreated or poorly treated domestic sewage. It is therefore likely that the reproductie performance of fish in the whole of the Napoleon Gulf could be endangered by endocrine disrupting estrogens.
5.4 17β-estradiol in water samples
The results show higher concentrations of 17β-estradiol in all water samples from Napoleon Gulf compared to estrone, and 17α-ethynylestradiol. Unexpectedly, samples from Kirinya; a study site presumed to hae no human actiities and therefore least polluted with human and animal wastes equally had significant amounts of estradiol. As it was the case with estrone, this finding implies that the whole of the Napoleon Gulf is heaily polluted with untreated or poorly treated human and animal wastewater. Like estrone, 17β-estradiol is a natural hormone excreted by both humans and animals through feces and urine and can only end up in a lake water enironment through discharge of untreated or poorly treated sewage into the lake (Swart & Pool, 2007). Although 17β-estradiol concentrations of up to 7000 ng/L were detected in water samples in the United Kingdom (Routledge et al., 1998), the concentrations of 17β-estradiol in water samples from Napoleon Gulf are higher than those of other studies. The concentration of 17β-estradiol in sewage effluent ranged from 2.7 to 48 ng/L in the United Kingdom, 3 ng/L in Germany, 64 ng/L in Canada, and 55 ng/L in Japan (Swart & Pool, 2007). Thus, the obtained high concentrations of estradiol in water samples could be greatly endangering the life of fish in the Napoleon Gulf of Lake Victoria.
5.5 17α-ethynylestradiol in water samples
Albeit in -arying concentrations, 17α-ethinylestradiol detected in all water samples from the Napoleon Gulf of Lake Victoria ranged from 45 to 360 ng/L. The detection of great amounts of this synthetic hormone in water samples compared to estrone implies that a large number of women in and around the study area use contraceptie pills containing 17α-ethynylestradiol and in turn excrete large quantities through urine and feces into the lake through sewage discharges. This finding could partly be explained by the promotional messages on different media houses about the use of synthetic birth control pills containing 17α-ethynylestradiol in family planning. This obseration is supported by Huang et al. (2013) who reported that use of birth control pills containing the synthetic 17α-ethynylestradiol and in turn its excretion in feces and urine is one of the major sources of synthetic estrogens in the -arious enironmental matrices. Howeer, the concentrations of 17α-ethynylestradiol in water samples from the study areas disagree with those of similar preious studies the fact that higher concentrations of the synthetic estrogen were determined in this study. According to Ying et al. (2009), low concentrations of 17α-ethynylestradiol were detected in water samples in the range of <LOD to 42 ng/L in Italy. In addition, Singhal et al. (2009) reported considerably lower concentrations of 17α-ethynylestradiol in water samples (0.1 to 5.1 ng/L) in New Zealand.
5.6 Estrogens in waste water samples from Lubigi channel, Kampala
In comparison, the study found higher concentrations of estrogens in water samples from Lubigi Channel in Kampala than in water samples from the Napoleon Gulf of Lake Victoria with exception of 17α-ethynylestradiol. This finding suggests that the channel drains mainly untreated wastewater generated in the city. This obseration is in agreement with other reported studies (Matagi, 2002; Naigaga et al., 2011; Swart & Pool, 2007). According to Matagi (2002), most of the pollutants from anthropogenic actiities in Kampala are discharged into Lake Victoria through different channels that drain the city. Furthermore, Naigaga et al. (2011) reported that most residents in Kampala central diision and other diisions are not connected to the national sewerage system especially households in the suburbs of the city. They further reported that the current sewerage system in Kampala city seres only a small fraction of the population and only 10% of all sewage generated in the city gets treated. This implies that oer 90% of sewage generated in Kampala remains untreated and is released into -arious drainage channels and ultimately Lake Victoria. The low income status and inability of most residents to get connected to the sewerage system exacerbates the problem. Consequently, wastes generated from different diisions are directly discharge into the -arious drainage channels including Lubigi Channel. This finding conforms to those of Swart and Pool (2007) who reported that humans and animals excrete hormones in wastes from their bodies that end up into the water enironment through sewage discharge. In contrast, the low concentration of 17α-ethynylestradiol in water samples from Kampala also suggests that a few of the residents use birth control pills containing 17α-ethynylestradiol.
5.7 Fish health and Production indices
From the results of estrogen analysis in water samples for estrogens, it might be concluded that the concentrations of estrogens had significant impact on the reproductie performance of fish. This conclusion is supported by Wright-Walter and Volz (2009) who reported that less than 1 ng/L of 17α-ethynylestradiol caused feminization of male fishes compared to 4 ng/L of 17β-estradiol. The GSI is often applied as an endpoint of endocrine disruption because a reduced GSI usually points to a reduction in gonad mass. A reduction in testes mass leading to reduced GSI has been described as an estrogenic effect (Mlambo et al., 2009). Howeer, the oerall high GSI ratios for male fish and female fish obtained in this study imply that EDEs had insignificant effect on the testicular tissue as well as oaries. According to Holland (2003), the possible effects of EDEs on the testis include degeneration of testicular tissue and generalized organ atrophy whereas the presence of atretic follicles and degeneratie oocytes are the common effects of EDEs in female fish and yet these histopathological changes were similarly obsered in this study. This finding could partly be attributed to drawback of this criterion where it is quite challenging to meet fish species that are synchronous and the interindi-idual -ariation in oarian weight during the spawning cycle. In addition, there is a problem of assigning causes of death/fatality in studies related to water due to the complexity of enironmental contaminants with similar effects on different biological and biochemical endpoints (Leino et al., 2005). Furthermore, the GSI reference ranges used in this study were determined experimentally on fish liing in a different enironment. This could explain the recorded disparities between the histopathological changes and GSI findings. Thus, a more thorough inestigation of the potential effects of EDEs on reproductie organs of fish would require conducting an experimental study rather than a crosssectional study.
The high HSI ratios obtained in this study are indicatie of normal liers of fish used in this study. The results show that female fish hae higher mean HSI -alue (2.67 + 0.21 %) than male fish (2.34 + 0.24 %). This implies that female fish are less affected by the chemical pollutants in water than males. Howeer, the HSI reference ranges used in this study were determined experimentally on fish liing in a different enironment and this could explain the obsered disparities. The lier is a major organ of metabolism and a reduction in fish lier mass would indicate effects of numerous chemical pollutants in water. According to Srijunngam and Wattanasirmkit (2001), this finding further suggests that the liers of O. niloticus were able to store enough ATP, an indicator of absence of toxicological effects of pollutants including estrogens on the lier of fish. Similarly, it is difficult to assign causes in studies related to water due to the complexity of enironmental contaminants with similar potential effects on different biological and biochemical endpoints (Leino et al., 2005). Howeer, the GSI and HSI results contradict the obsered histopathological changes. This therefore calls for an experimental study to highlight the obsered disparities.
5.8 Histopathological findings
During oariogenesis, oocytes undergo different maturation stages including preitellogenic oocytes and mature -itellogenic oocytes (Mlambo et al., 2009). Similarly, during spermatogenesis, spermatogonia undergo numerous mitotic diisions forming primary spermatocytes, secondary spermatocytes, spermatids and spermatozoa; resulting in production of cysts surrounded by sertoli cells. Each lobule of the testes contains numerous cysts. Regularly organized cysts with all stages of spermatogenesis are indicatie of testes in a healthy reproductie condition (Dumitrescu et al., 2010). Howeer, histopathological changes such as presence of inflammatory cells, focal areas of necrosis, tissue congestion and degeneration as well as siderotic nodules in the spleen and the lier are suggestie of the possible effects of different types of water chemical pollutants (Srijunngam & Wattanasirmkit, 2001).
Conersely, the presence of necrotic and fibrotic areas in the testicular tissue was indicatie of seere degeneratie changes in organs of male fish whereas the presence of atretic follicles in the oaries of female fish was the common histopathological finding . Since higher concentrations of estrogens (about 9 times and more the lowdose effect of estrogens - 1 ng/L) were detected in water samples, histopathological changes obsered in -arious tissues of fish could be attributed to EDEs. Additionally, histopathological changes obsered in gonads did show a clear interruption of the normal pattern of oocytes maturation normally indicatie of inhibition of steroidogenesis. McKinlay et al. (2008) reported that inhibition of oocyte maturation is linked to inhibition of gonadotrophic hormone secretion, thus, EDEs presence in water could partly explain the occurrence of the obsered histopathological changes in fish oaries.
It was further reported that the presence of atretic follicles in the oarian tissue is indicatie of the effects of endocrine disrupting chemicals including EDEs on oarian deelopment and spawning (Mlambo et al., 2009). Howeer, it is difficult to conclude that the obsered histopathological changes were a direct result of EDEs exposure since most chemical pollutants cause similar toxicological effects in different tissues of O. niloticus (Leino et al., 2005). In this study, most of the fish used in histopathological examination were caught from areas that were distant from the water sampling points. Thus, a more thorough inestigation of the potential effects of EDEs on the histological picture of different fish tissues and at different deelopmental stages of fish is recommended.
CHAPTER SIX
6.0 CONCLUSIONS AND RECOMMENDATIONS
6.1 Conclusions
- This was the first study to report on the presence of endocrine disrupting estrogens in Lake Victoria waters in Uganda.
- Water samples from Napoleon Gulf of Lake Victoria Uganda contain high leels of endocrine disrupting estrogens (about 9 times and more the estrogen lowdose effect; 1 ng/L).
- Tissue samples from fish caught in the study area had distinct histopathological changes possibly due to exposure to different chemical pollutants including EDEs in water.
- GSI and HSI results show no significant difference from the reiewed normal ranges but contradict with the obsered histopathological changes.
- The physicochemical analysis of water was suggestie of poor quality of water.
- Comparatie analysis of estrogen in water samples from Lubigi channel in Kampala show much higher leels of estrogens than water samples from Napoleon Gulf, Lake Victoria except for 17α- ethynylestradiol.
- Histopathological changes obsered in the testicular and oarian tissues of fish indicate possible impacts of EDEs on fish reproduction and decline in fish catches.
6.2 Recommendations
On the basis of the current findings it is recommended that;
- Since humans would generally take pills and consume -arious products including plant products in addition to endogenous estrogens, it would be better to assay domestic water to determine the leels of exposure to enironmental EDEs.
- Indiidual and combinational effects of endocrine disrupting estrogens need inestigation to elucidate their synergistic and antagonistic effects on different fish species.
- Methods that remoe estrogens from wastewater such as actiated sludge wastewater treatment plants should be incorporated into treatment processes to control discharge of estrogenic wastes into Lake Victoria to safeguard fish, domestic animals, and humans from the health hazards of these estrogens.
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APPENDICES
APPENDIX I: ELISA absorbance readings of estrogens in the sample, standards and controls using a spectrophotometer (Gene 5) at 450 nm
A: Total estrogens
Abbildung in dieser Leseprobe nicht enthalten
B: Estrone
Abbildung in dieser Leseprobe nicht enthalten
C: 17 β-estradiol
Abbildung in dieser Leseprobe nicht enthalten
D: 17 α-ethynylestradiol
Abbildung in dieser Leseprobe nicht enthalten
Key: Wells with grey background= standards; wells with a pale yellow background-= control; wells with clear background= samples & wells of the Column 12 (Dark background) were not used.
APPENDIX II: Measurement and collection of organs from fish in pictures
Abbildung in dieser Leseprobe nicht enthalten
Key: A: the fish oaries; B: Measurement of total length of fish; C: display of dissected organs from fish; D: buckets for safe transport of fish; E: taking the total body weight of fish; and F: transferring the collected specimen into a fixatie solution.
APPENDIX III: Different wastewater sample collection sites in pictures
Abbildung in dieser Leseprobe nicht enthalten
Key: A: Marine and Agroprocessing plant at Kirinya; B: a drainage channel in Wanyange; C: Leather and Tanning factory at Kirinya; and D: a heaily polluted sampling point at one of the shores.
APPENDIX IV: Collection of fish samples and laboratory analysis of estrogens in water samples at Napoleon Gulf of Lake Victoria and Makerere Uniersity respectiely in pictures.
Abbildung in dieser Leseprobe nicht enthalten
Key: H: catching fish from the study area; I: agricultural actiities taking place adjacent to the lake; J: taking pH measurement in the Laboratory; K: extraction of estrogens from the water samples.
APPENDIX V: Standard cures for ELISAs
Typical standard cures for total estrogens, estrone, 17β-estradiol, and 17α-ethynylestradiol are illustrated in figures 4-5. The sensitiities of the ELISA kits were determined by the supplier, Japan Eniro Chemicals Ltd.
Abbildung in dieser Leseprobe nicht enthalten
Figures 4A & 4B: The ELISA for estrone has a detection range between 0.05- 5.0 µg/L (Figure 4A). The 17β-estradiol ELISA has a range between 0.05- 1.0 µg/L (Figure 4B).
Abbildung in dieser Leseprobe nicht enthalten
Figure 5C & 5D: The 17α-ethynylestradiol ELISA has a detection range between 0.05 and 3.0 µg/L (Figure 5C). The ELISA for total estrogens has a detection range of 0.05 and 3.0 µg/L (Figure 5D).
APPENDIX VI: Water sampling sites in Napoleon Gulf of Lake Victoria and some areas in Kampala Uganda.
Abbildung in dieser Leseprobe nicht enthalten
APPENDIX VII: Some of the histopathological changes obsered in different tissues of O. niloticus caught from Napoleon Gulf of Lake Victoria, Uganda .
Abbildung in dieser Leseprobe nicht enthalten
APPENDIX VII: Data collection forms
MAKERERE UNIVERSITY COLLEGE OF VETERINARY MEDICINE, ANIMAL RESOURCES AND BIOSECURITY
Re: Analysis of endocrine disrupting estrogens in water and pathological changes in tissues of Oreochromis niloticus Lake Victoria, around Jinja municipality:
Data collection sheet for recording results of water quality parameters
Study site ID No….
A: Measurement of physicalchemical parameters
Abbildung in dieser Leseprobe nicht enthalten
MAKERERE UNIVERSITY COLLEGE OF VETERINARY MEDICINE, ANIMAL RESOURCES AND BIOSECURITY
Re: Analysis of endocrine disrupting estrogens in water and pathological changes in tissues of Oreochromis niloticus Lake Victoria, around Jinja municipality:
Data collection sheet for water samples
Study site ID No….
B: Collection of water samples for ELISA
Abbildung in dieser Leseprobe nicht enthalten
Obserations …
MAKERERE UNIVERSITY COLLEGE OF VETERINARY MEDICINE, ANIMAL RESOURCES AND BIOSECURITY
Re: Analysis of endocrine disrupting estrogens in water and pathological changes in tissues of Oreochromis niloticus Lake Victoria, around Jinja municipality:
Data collection sheet for fish samples
Study site ID No….
C: Collection of histological samples
Abbildung in dieser Leseprobe nicht enthalten
- Quote paper
- Saphan Muzoora (Author), 2013, Analysis of Endocrine Disrupting Estrogens in Water and Pathological Changes in Tissues of Oreochromis Niloticus from Lake Victoria, Uganda, Munich, GRIN Verlag, https://www.grin.com/document/377655
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