Claire Michelle Volschenk

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How to cite this thesis

Surname, Initial(s). (2012). Title of the thesis or dissertation (Doctoral Thesis / Master’s Dissertation). Johannesburg: University of Johannesburg. Available from: http://hdl.handle.net/102000/0002 (Accessed: 22 August 2017).

TOXICITY TESTING FOR THE ESTABLISHMENT OF LC50 AND EC50 VALUES FOR TECHNICAL GRADE DDT, IN CONJUNCTION WITH BIOMARKERS AND BIO-ACCUMULATION IN SYNODONTIS ZAMBEZENSIS.

CLAIRE MICHELLE VOLSCHENK

THESIS

SUBMITTED IN FULFILMENT OF THE REQUIREMENTS OF THE DEGREE PHILOSOPHIAE DOCTOR

ZOOLOGY

IN THE

FACULTY OF SCIENCE

AT THE

UNIVERSITY OF JOHANNESBURG

SUPERVISOR: PROF. RICHARD GREENFIELD

CO-SUPERVISOR: PROF. JOHANNES H.J VAN VUREN

NOVEMBER 2018

TABLE OF CONTENTS

LIST OF FIGURES ...... vii LIST OF TABLES ...... ix LIST OF ABBREVIATIONS ...... x ACKNOWLEDGEMENTS ...... xvi SUMMARY ...... xviii Chapter 1: General introduction ...... 1 1. General introduction ...... 1 1.1 Background...... 2 1.2 Biological markers and selected indicator species ...... 3 1.3 Motivation for study ...... 4 1.4 Hypotheses, aims and objectives ...... 5 1.4.1 Hypotheses ...... 6 1.4.2 Aims ...... 6 1.4.3 Objectives ...... 7 1.5 Layout of thesis ...... 7 Chapter 1: General introduction ...... 7 Chapter 2: Literature review ...... 8 Chapter 3: Focus species ...... 8 Chapter 4: Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852) ...... 8 Chapter 5: Baseline bio-accumulation concentrations and resulting oxidative stress in Synodontis zambezensis after an acute laboratory exposure to 4, 4’- DDT ...... 9 Chapter 6: Biomarkers of exposure and energetics ...... 9 Chapter 7: Ribonucleic acid (RNA) as a biomarker ...... 9

Chapter 8: LC50 exposures and species sensitivity distributions ...... 9 Chapter 9: Conclusions and recommendations ...... 9 Chapter 10: References ...... 9 Appendix A: Raw data ...... 9 Appendix B: Published papers ...... 10 Appendix C: Conference contributions ...... 10 1.6 Anticipated contribution ...... 10

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Chapter 2: Literature review ...... 11 2. Introduction ...... 11 2.1 4,4’-DDT ...... 12 2.1.1 Historical use of 4,4’-DDT ...... 12 2.1.2 The use of 4,4’-DDT in South Africa ...... 14 2.1.3 Intended target of 4,4’-DDT ...... 15 2.1.4 The fate of 4,4’-DDT in the environment ...... 16 2.2 Bio-accumulation ...... 18 2.2.1 Absorption and regulation ...... 19 2.3 Effects of 4,4’-DDT on fish and offspring ...... 20 2.3.1 Developmental toxicity ...... 21 2.4 Toxicity testing framework ...... 22 2.4.1 Acute exposure...... 22 2.4.2 Acute fish toxicity test (OECD TG, 203) ...... 23 2.5 Biomarkers...... 23 2.5.1 Biomarkers of exposure ...... 24 Cytochrome P450 (CYP450) ...... 24 Acetylcholinesterase (AChE)...... 25 Cellular energy allocation (CEA) ...... 25 2.5.2 Biomarkers of effect ...... 26 Oxidative stress biomarkers ...... 26 Superoxide dismutase (SOD) ...... 26 Catalase (CAT) ...... 26 Malondialdehyde (MDA) ...... 27 Protein carbonyls (PC) ...... 27 Reduced glutathione (GSH) ...... 27 2.5.3 RNA as a biomarker ...... 28 Chapter 3: Focus species ...... 29 3. Introduction ...... 30 3.1 Taxonomy ...... 31 3.2 External anatomy ...... 32 3.3 Distribution ...... 33 3.4 Feeding habits ...... 34

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3.5 Artificial spawning and its importance ...... 34 Chapter 4 : Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852) ...... 36 Abstract ...... 37 4. Introduction ...... 37 4.1 Materials and methods ...... 38 4.1.1 Water quality preferences ...... 39 4.1.2 Artificial breeding setup ...... 39 4.1.3 Aquaspawn® Injection, egg stripping and fertilisation ...... 39 4.1.4 Egg tumbling and holding tank ...... 40 4.1.5 Larvae, fry and fingerling maintenance ...... 41 4.1.6 Microscopic observation of egg development ...... 42 4.2 Results and discussion ...... 42 4.2.1 Breeding ...... 42 4.2.2 Embryonic development ...... 42 4.3 Conclusions and recommendations ...... 44 Chapter 5 : Baseline bio-accumulation concentrations and resulting oxidative stress in Synodontis zambezensis after an acute laboratory exposure to 4,4’- DDT ...... 45 Abstract ...... 46 5. Introduction ...... 47 5.1 Materials and methods ...... 49 5.1.1 Fish collection ...... 49 5.1.2 Acclimation ...... 49 5.1.3 Exposure concentrations ...... 50 5.1.4 Exposure system and transfer of fish ...... 50 5.1.5 Water quality ...... 51 5.1.6 Tissue sample collection ...... 51 5.1.7 Water sample collection ...... 51 5.1.8 DDT analysis of tissue and water ...... 52 Tissue extraction ...... 52 Water extraction ...... 53 Gas chromatography-mass spectrometry ...... 53 Quality control and quality assurance ...... 54

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5.1.9 Biomarker analysis of muscle and gill tissue ...... 55 Superoxide dismutase ...... 55 Catalase ...... 55 Malondialdehyde ...... 56 Protein carbonyl ...... 56 Reduced glutathione ...... 56 5.1.10 Statistics ...... 56 5.2. Results ...... 57 5.2.1 Biometric data...... 57 5.2.2 Water quality ...... 57 5.2.3 Levels of 4,4’-DDT in Synodontis zambezensis...... 58 5.2.4 Levels of 4,4’-DDT in water samples ...... 60 5.2.5 Oxidative stress biomarkers ...... 61 5.2.6 Relationship between biomarkers and bio-accumulation ...... 63 5.3 Discussion ...... 66 5.3.1 Water quality ...... 66 5.3.2 Levels of 4,4’-DDT in Synodontis zambezensis...... 66 5.3.3 Levels of 4,4’-DDT in water samples ...... 68 5.3.4 Oxidative stress biomarkers and the link to bio-accumulation ...... 69 5.4 Conclusions and recommendations ...... 71 5.6 Acknowledgements ...... 72 Chapter 6: Biomarkers of exposure and energetics ...... 73 6. Introduction ...... 74 6.1 Materials and methods ...... 76 6.1.1 Fish tissue collection ...... 76 6.1.2 Biomarker protocols ...... 76 Biomarkers of exposure ...... 76 Acetylcholinesterase ...... 76 Cytochrome P450 ...... 77 Biomarkers of energetics ...... 77 Cellular energy allocation ...... 77 Energy availability ...... 78 Energy consumption ...... 79

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6.1.3 Statistical analysis ...... 80 6.2 Results ...... 80 6.2.1 Biomarkers of exposure and energetics vs. 4,4’-DDT ...... 80 6.3 Discussion ...... 86 6.3.1 Biomarkers of exposure and energetics vs. 4,4’-DDT ...... 86 6.4 Conclusions and recommendations ...... 88 Chapter 7: Ribonucleic acid (RNA) as a biomarker ...... 89 7. Introduction ...... 90 7.1 Materials and methods ...... 98 7.1.1 Fish tissue collection ...... 98 7.1.2 RNA isolation reference ...... 98 7.1.3 cDNA synthesis ...... 100 7.1.4 Primer design protocol ...... 101 7.1.5 General PCR – Primer specificity ...... 102 7.1.6 Sub cloning ...... 103 Petri dish preparation ...... 103 TOPO® Cloning reaction ...... 103 Rapid One Shot® chemical transformation protocol ...... 103 Analysis of transformants by PCR ...... 103 Plasmid purification ...... 105 7.1.6 qRT - PCR ...... 105 The dilution method ...... 105 Primer efficiency calculations ...... 106 7.2 Results and discussion ...... 106 7.3.1 Primer efficiency ...... 109 7.4 Conclusions and recommendations ...... 110

Chapter 8: LC50 exposures and species sensitivity distributions ...... 112 8. Introduction ...... 113 8.1 Materials and methods ...... 116 8.1.1 Fish species ...... 116 8.1.2 Acclimation ...... 116 8.1.3 Experimental setup ...... 116 8.1.4 Water quality ...... 118

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8.1.5 Statistics ...... 118 8.2 Results ...... 120 8.2.1 Water quality ...... 120

8.2.2 LC50 Transformative Probit analysis ...... 121 8.3 Discussion ...... 123 8.3.1 Water quality ...... 123

8.3.2 LC50 Transformative Probit analysis ...... 123 8.4 Conclusions and recommendations ...... 124 Chapter 9: Conclusions and recommendations ...... 125 9. Conclusions and recommendations ...... 126 9.1 Laboratory based exposures ...... 126 9.1.1 Recommendations ...... 128 9.2 Biomarker analysis ...... 128 9.2.1 Recommendations ...... 130 9.3 Artificial breeding of Synodontis zambezensis ...... 130 Chapter 10: References ...... 132 Appendix A: Raw Data ...... 157 Appendix B: Published Manuscript ...... 172 Appendix C: Conference Contributions ...... 175

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LIST OF FIGURES Figure 2. 1: Degradation of 4,4’-DDT to form its metabolites, 4,4’-DDE (removal of HCl) and 4,4’-DDD (through reductive dechlorination) (Li et al., 2010)...... 12 Figure 2. 2: Graphical representation of the primary structure of voltage-gated sodium channel with α and β units (Catterall, 2000)...... 15 Figure 2. 3: Illustration of the “Grasshopper Effect” also known as global distillation. Adapted from Bhardwaj et al.(2018)...... 18 Figure 2. 4: Trade-off between response sensitivity and ecological relevance over time at the different levels of biological organisation. Adapted from Adams, (2001). 21 Figure 3. 1: Image of Synodontis sp. adapted from Skelton (2001)...... 32 Figure 3. 2: External sex differentiation of female (A) and male (B) Synodontis zambezensis indicating the anal orifice and the genital papillae...... 33 Figure 3. 3: Distribution of Synodontis zambezensis within Africa. Grey highlighting indicates IUCN distribution data and green highlighted countries indicate published data sets on this species (Bills et al., 2010)...... 34 Figure 4. 1: Adult fish holding tank (100 L) setup with false bottom, sponge filter, 25 W heater and PVC piping...... 39 Figure 4. 2: Artificial fertilisation process; A) stripped eggs placed into dry glass Petri dish, B) semen from testes added to eggs and mixed with feather C) water from hatching tank added and mixed with sperm and eggs using a feather...... 40 Figure 4. 3: Do it yourself (DIY) Egg Tumbler in a 3 L beaker with Fungus Clear treated fish medium...... 41 Figure 5. 1: Biomarkers of Oxidative stress within the liver and the gills of Synodontis zambezensis after an acute exposure to 4,4’-DDT. Superoxide dismutase (Figure 1A), PC (Figure 1B), CAT (Figure 1C), GSH (Figure 1D) and MDA (Figure 1E). Significant differences (p≤0.05) between tissues within the same biomarker are indicated by the presence of capital letter suberscripts. Significant diffierences (p≤0.05) within the same tissue (liver or gills) of the same biomarker are indicated by common lower case letter suberscripts...... 62 Figure 5. 2: RDA triplot showing resulting oxidative stress biomarker responses in Synodontis zambezensis following an acute exposure to varying concentrations of 4,4’-DDT. The resulting oxidative stress biomarkers are constrained in this ordination by 4,4’-DDT and its degradation products, 4,4’-DDE and 4,4’-DDD...... 65 Figure 6. 1: Biomarkers of exposure (mean ±SD) within the liver (A), AChE in the liver and muscle tissue (B), energetics (C) and Ea and its components (D) in Synodontis zambezensis after an acute exposure to various concentrations of 4,4’- DDT. Significant differences (p≤0.05) are indicated by the presence of common superscripts (capital letters or lower case letters). Resulting biomarker concentrations have been log transformed in order for them to be illustrated together on the corresponding graphs...... 84 Figure 6. 2: An RDA triplot showing resulting biomarkers of exposure and energetics in Synodontis zambezensis following an acute exposure to 4,4’-DDT...... 86

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Figure 7. 1: Polymerase Chain Reaction steps and components (Chou et al., 1992)...... 93 Figure 7. 2: Figure showing the presence of both white (successful) and blue (unsuccessful) bacterial colonies (Padmanabhan et al., 2011)...... 94 Figure 7. 3: Resulting separation after the centrifugation during TRIzol RNA extraction (Hummon et al., 2007)...... 99 Figure 7. 4: The blue-ring filter tube supplied in the Nucleospin®RNA isolation kit (Macherey-Nagel, 2015)...... 99 Figure 7. 5: Gene alignment example (Larkin et al., 2007), with an asterisk (*) indicating a single fully conserved residue, a colon (:) indicating conserved mutations, a full stop (.) indicating semiconservative mutations and an empty space ( ) indicating non-conservative mutations...... 101 Figure 7. 6: Figure showing the both unsuccessful (A) colony culture in LB medium as well as successful (B) colony growth in LB medium after 16 hour incubation period...... 104 Figure 7. 7: Amplification of PCR products for ER (B) and β-actin (C) in muscle (1), liver (2), kidney (3) and gonadal (4) tissue of Synodontis zambezensis...... 107 Figure 7. 8: Amplification of PCR products for the transferrin gene in muscle (1), liver (2), kidney (3) and gonadal (4) tissue of Synodontis zambezensis...... 108 Figure 7. 9: Regression lines indicating the efficiency of the β-actin forward and reverse primers within the extracted tissues (Liver, kidney, testes and gonads) of Synodontis zambezensis...... 110 Figure 8. 1: Oreochromis mossambicus, picture taken from Cambray and Swarts (2007). Photo credit Dr D. Tweddle...... 114 Figure 8. 2: A male (above) and female (below) Trinidadian guppy (Deacon et al., 2015)...... 115 Figure 8. 3: Set up of 500 mL beakers. Both controls and all exposure concentrations ranging from 0.6 ng/mL to 15.4 ng/mL were set up in duplicates in an environmental room at 25˚C with a 12h/12h day/night photoperiod...... 117 Figure 8. 4: Linear regression graphs from 96-hour exposure to 4,4’-DDT of juvenile S. zambezensis (A), P. reticulata (B) and O. mossambicus (C) as well as the distribution of sensitivity between these three species (D)...... 122

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LIST OF TABLES Table 4 1: Time after fertilisation, description of developmental stages and photographs of Synodontis zambezensis bred in captivity...... 43 Table 5. 1: Biometric data (mean ± SD) and range of Synodontis zambezensis exposed to 4,4’-DDT during an acute exposure...... 57 Table 5. 2: In situ water quality parameters (mean ± SD) and range readings taken of the exposure tanks during the course of the 96-hour 4,4’-DDT exposure experiment. pH, dissolved oxygen (DO), oxygen percentage (O2%), electrical conductivity (EC), temperature (˚C) and total dissolved solids (TDS)...... 58 Table 5. 3: Levels of accumulation in S. zambezensis muscle tissue and concentrations in water samples of the 4,4’-DDT degradation products (mean ± SD) and range after an acute laboratory exposure...... 59 Table 5. 4: Levels of accumulation of the 4,4’-DDT degradation products in S. zambezensis muscle tissue and extracted from the exposure tank water as well as percentage 4,4’-DDT accumulated in S. zambezensis. The 4,4’-DDT used was of 98.7% purity purchased from Sigma-Aldrich PESTANAL® as an Analytical Standard...... 60 Table 5. 5: Tabulated correlation coefficients (factor or component loadings) between the resulting oxidative stress biomarkers (SOD,MDA, PC, CAT and GSH) in both the liver and gill tissue of Synodontis zambezensis in relation to 4,4’-DDT and its degradation products 4,4’-DDE and 4,4’-DDT...... 65 Table 6. 1: Resulting concentrations of AChE and CYP450 (Mean ± SD) and range in the liver tissue of Synodontis zambezensis after an acute exposure to 4,4’-DDT. Significant differences (p≤0.05) per column are indicated by common superscripts in the form of capital or lower case letters...... 81 Table 6. 2: Resulting concentrations of AChE, CEA, Ec, Ea, glucose, lipids and proteins (mean ± SD) and range in the muscle tissue of Synodontis zambezensis after an acute exposure to 4,4’-DDT. Significant differences (p≤0.05) per column are indicated by common superscripts in the form of capital or lower case letters...... 83 Table 7. 1: Forward and reverse primer sets compiled for the amplification of β-actin, transferrin and oestrogen receptor (ER) genes...... 106 Table 7. 2: Slope of regression lines and resulting efficiencies of ER, Transferrin and β-actin forward and reverse primers for successfully sequenced corresponding genes in Synodontis zambezensis...... 109 Table 8. 1: Table showing the transformation values of percentage mortality into probability units (Probits) (Finney, 1947)...... 119 Table 8. 2: In situ water quality parameters (mean ± SD) and range readings taken of the exposure beakers during the course of the 96-hour 4,4’-DDT LC50 exposure experiment. Electrical conductivity (EC), pH, oxygen percentage (O2%) and temperature (˚C)...... 120 Table 8. 3: Species, regression equation and the solutions for x as described in the statistics section. The resulting LC50 values indicated the sensitivities of each of the exposed species...... 123

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LIST OF ABBREVIATIONS %: Percent ...... 12 ˚: Degrees ...... 64 ˚C/minute: Degrees Celsius per minute ...... 52 ˚C: Degrees Celsius ...... 38 <: Less than...... 84 ±: Plus/minus ...... 41 ≤: Less than or equal to ...... 45 ≥: Greater than or equal to ...... 51 µg/L: Micrograms per litre...... 51 µg/mg protein: Micrograms per milligram protein ...... 63 µL: Microlitres ...... 74

µmol H2O2/min/mg protein: Micromoles Hydrogen Peroxide per minute per milligram protein ...... 63 µS/cm: Microsiemens per centimetre ...... 57 3’: Three Prime ...... 89 5’: Five Prime ...... 89 A: Adenine ...... 87 abs/min/mg: Absorbance per minute per milligram protein ...... 78 ACh: Acetylcholine ...... 24 AChE: Acetylecholinesterase ...... 5 ANOVA: Analysis of Variance ...... 55 ATP: Adenosine Triphosphate ...... 73 BCF: Bio-concentration Factor ...... 18 BLAST: Basic Local Alignment Search Tool ...... 106 bp: Base pairs ...... 92 BSA: Bovine Serum Albumin ...... 54 BSS: Buffered Sustrate Solution ...... 77 C. gariepinus: Clarias gariepinus ...... 2 C: Cytosine ...... 87

C6H5Cl: Chlorobenzene ...... 12

CaCl2: Calcium Chloride ...... 39 CAT: Catalase ...... 5

CCl3CHO: Chloral ...... 12 cDNA: Complimentary DNA ...... 9

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CEA: Cellular Energy Allocation ...... 5 CHg H: Choriogenin H ...... 93 CHg L: Choriogenin L ...... 93 Chgs: Choriogenins ...... 94 cm: Centimetres ...... 38 CYP: Cytochrome P ...... 93 CYP1A: Cytocrome 1A ...... 27 CYP450: Cytochrome P450 ...... 5 D.O: Dissolved Oxygen ...... 56 DCM: Dichloromethane ...... 51 DDT: 1,1'-(2,2,2-Trichloro-1,1-ethanediyl)bis(4-chlorobenzene) ...... 2 DDW: Double Distilled Water ...... 99 DH5α: Douglas Hanahan 5 alpha ...... 92 DNA: Deoxyribonucleic Acid ...... 25 dNTPs: deoxy-ribonucleotide triphosphates ...... 88 DOH: Department of Health ...... 14 E. coli: Escherichia coli ...... 91 Ea: Energy availability ...... 5 EB: Elution Buffer ...... 102 EC: Electrical Conductivity ...... 57 Ec: Energy consumption ...... 5 EDC: Endocrine Disrupting Chemical ...... 94 EDTA: Tris-Borate-Ethylnenediaminetetraacetic acid ...... 92 EI-MS: Electron Ionization-Mass Spectrometry ...... 53 EPA: Environmental Protection Agency ...... 52 ER: Oestrogen Receptor ...... 21 ERE: Oestrogen Responsive Elements ...... 94 ETS: Electron Transport System ...... 25 Ex Taq: Thermus aquaticus ...... 88 FASMAC: Food Assessment and Management Centre ...... 105 FET: Fish Embryo Test ...... 121 g.: G-force ...... 54 g: Grams ...... 39 G: Guanine ...... 87 g6pd: Glucose-6-phosphate dehydrogenase ...... 93

xi | P a g e gapdh: Glyceraldehyde 3-phosphate dehydrogenase ...... 93 GC: Guanine+Cytosine ...... 89 GC-MS: Gas Chromatography-Mass Spectrophotometry ...... 51 gDNA: Genomic Deoxyribonucleic Acid ...... 98 GHB: General Homogenising Buffer ...... 54 GOI: Gene of Interest ...... 95 GPC: Gel Permeation Chromatography ...... 51 GSH: Reduced Glutathione ...... 5 GSSG: Glutathione Disulphide ...... 69

H2SO4: Sulphuric Acid ...... 12 HK: Housekeeping ...... 93 hr: Hour ...... 38 hrs: Hours ...... 38 HSP: Heat Shock Proteins ...... 85 INT: 3-(4-Iodophenyl)-2-(4-nitrophenyl)-5-phenyl-2H-tetrazol-3-ium chloride ...... 77 IRS: Indoor Residual Spraying ...... 14 IUCN: International Union for Conservation of Nature...... 32 IUPAC: International Union of Pure and Applied Chemistry ...... 12 J/g: Joules per gram ...... 79 kbp: Kilo base-pairs ...... 92 KCl: Potassium Chloride ...... 39

KMnO4: Potassium Permanganate ...... 54

Kow: Log octanol-water partition coefficient...... 18 L: Litres ...... 38 LB: Lysogeny Broth ...... 91 LC: Lowest Concentration ...... 110

LC50: Lethal Concentration 50 ...... 4 LD: Lowest Dose ...... 110 m/z: Mass to charge ratio ...... 53 M: Molar ...... 54 MDA: Malondialdehyde ...... 5 MDB: Membrane Desalting Buffer ...... 97 mg/L: Milligram per litre ...... 56

MgSO4: Magnesium Sulphate ...... 39 min: Minute ...... 75

xii | P a g e mins: Minutes ...... 75 mL: Millilitres ...... 39 mm: Millimetres ...... 38 mRNA: Mitochondrial RNA ...... 87 MSDS: Material Safety Data Sheet ...... 10 NaCl: Sodium Chloride ...... 91 NADPH: Reduced nicotinamide-adenine dinucleotide phosphate ...... 75

NaHCO3: Sodium Bicarbonate ...... 39 NaOH: Sodium Hydroxide ...... 91 NCBI: National Centre for Biotechnology Information ...... 98 ng SOD/mg protein: Nanograms SOD per milligram protein ...... 60 ng/g: Nanograms per gram ...... 45 ng/mL: Nanograms per millilitre ...... 49 ng/μL: Nano grams per microliter ...... 97 nM/mg protein: Nanomoles per milligram protein ...... 78 nm: Nanometres ...... 54 nmol carbonyl/mg protein: Nanomoles carbonyl per milligram protein ...... 62 nmol/mg protein: Nanomoles per milligram protein ...... 61 NOEC: No Observed Effect Concentration ...... 114 NRF: National Research Foundation...... 70 O. mossambicus: Oreochromis mossambicus ...... 2

O2%: Oxygen Percentage ...... 56 OCPs: Organochlorine Pesticides ...... 2 OECD: Organisation for Economic Cooperation and Development ...... 22 P. reticulata: Poecilia reticulata ...... 6 PAH’s: Polyaromatic Hydrocarbons ...... 23 PC: Protein Carbonyls ...... 5 PCB #77: 3,3'4,4'-tetrachlorobiphenyl ...... 52 PCB’s: Polychlorinated Biphenyls ...... 23 PCR: Polymerase Chain Reaction ...... 87 pg: Pico gram ...... 98 POPs: Persistent Organic Pesticides ...... 2 ppb: Part per billion ...... 113 ppm: Parts per million ...... 113 Probit: Probability Unit ...... 7

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PVC: Polyvinyl Chloride ...... 38 qRT-PCR: Quantitative real-time PCR ...... 95 r.p.m: Revolutions per minute ...... 76 R²: Coefficient of Determination ...... 118 RDA: Redundancy Analysis ...... 55 RNA: Ribonucleic Acid ...... 5 ROS: Reactive Oxygen Species ...... 25 rRNA: Ribosomal RNA ...... 87 S. nigromaculatus: Synodontis nigromaculatus ...... 41 S. petricola: Synodontis petricola ...... 36 S. zambezensis: Synodontis zambezensis ...... 2 s-Acetylthiocholine iodide: 2-(Acetylsulfanyl)-N,N,N-trimethylethanaminium iodide . 74 SANS: South African National Standard ...... 37 SD: Standard Deviation ...... 60 sec: Second ...... 54 SIM: Selected Ion Monitoring ...... 53 SOD: Superoxide Dismutase ...... 5 SSDs: Species Sensitivity Distributions ...... 6 T: Thymine ...... 87 TCmX: 2,4,5,6-tetrachloro-m-xylene ...... 51 TDS: Total Dissolved Solids ...... 50 TG: Target Gene ...... 101 Tm: Melting Temperature ...... 89 tRNA: Transfer RNA ...... 87 U: Uracil ...... 87 USEPA: United States Environmental Protection Agency ...... 13 VTG I: Vitellogenin I ...... 93 VTG II: Vitellogenin II ...... 93 VTGs: Vitellogenins ...... 94 WHO: World Health Organisation ...... 13 WWII: World War II………………………………………………………………………...13 WRC: Water Research Commission ...... 4 X-gal: 5-Bromo-4-chloro-1H-indol-3-yl β-D-galactopyranoside ...... 92 β-actin: Beta-actin ...... 93 μg: Micro gram ...... 98

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μL: Microliters ...... 52 μM: Micro moles ...... 99

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ACKNOWLEDGEMENTS I would like to sincerely thank the following people and organizations; I would not have been able to complete this study without the support of each and every one of you:

University of Johannesburg (Global Excellence Stature Scholarship), the National Research Foundation (NRF – UID: 102440) and Hokkaido University (UID 92424 Japan/ SA Bi-lateral Programme) for their financial support throughout this thesis.

Department of Chemistry at the University of Johannesburg: Dr Elize Smit, Dr Noh and Prof Meijboom for their support and use of equipment

Department of Zoology at the University of Johannesburg: Use of equipment in both the Ecotoxicology Laboratory and the environmental rooms in the aquarium

Veterinary Toxicology Laboratory and Professors at Hokkaido University, Japan Use of equipment and support in analysis of samples forming part of the bio- accumulation and genetics section of this thesis

Shimadzu South Africa Mr Harris and Dr Meyer for the use of equipment for analysis of samples and support in training on GC-MS

Colleagues in the Zoology Department and Ecotoxicology Laboratory To all of my colleagues, thank you for all of your support and advice throughout all the ups and downs of this thesis. You have all become my family during the last 6 years and I don’t think I could ever have done it without all of you. Special thanks must go specifically to Mrs Simone Dahms-Verster and Mr Gregg Jansen van Rensburg.

Dr Amina Nel My lab Mom – thank you for all of your support throughout my university career, all the way from first year. Your help and guidance throughout the past 9 years has been invaluable.

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Prof. J.H.J. van Vuren My co-supervisor – thank you for all of your support, professional advice and constant patience and guidance throughout this thesis.

Prof. R. Greenfield My supervisor – thank you for the words of advice and support through presentations and experiments. Thank you for all of the hours you have spent reading through this thesis and for making the past 4 years a great experience.

My Family Joanne and William Jones, you have done more than any parents would do for their children. Words cannot express how grateful I am for all of the support you have given me throughout the years of study, specifically over this last year.

Nina, you are the best big sister I could ever have asked for and even though you’re far away, your support was always just a message or a phone call away. I love you very much.

Bruce Edwards, thank you for being a positive, supportive part of my life and for always believing in me. I appreciate you so much and I cannot express how grateful I am to have you in my life. I love you lots Dad.

My late grandmother, Margaret-Ann Loots (1946-2018) I wish that you were here to celebrate this milestone with me. Thank you for being such a rock for me throughout my life and I look forward to seeing you again. Love you forever.

My Husband Mr Rikus Volschenk you are my best friend, my safe place and my biggest fan. Thank you for everything you have given up to allow me to follow my career dreams. You are my favourite person and I can’t wait to celebrate this milestone with you. I love you to the moon and back.

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SUMMARY The degradation of freshwater aquatic ecosystems is a problem occurring in countries throughout the world. This degradation can be attributed to the input of human waste products and wastewater resulting from agricultural and mining practices. Since South Africa is classified as a 3rd world country, the development and growth in the industrial and agricultural sectors far outweigh any remediation taking place in response to existing environmental damage. Understanding the effects of xenobiotics and pollutants on aquatic systems is an important step in improving the management of impacted areas.

Many pesticides are characteristically lipophilic, allowing them to accumulate within organisms occupying freshwater and marine systems. The organisms living within affected systems may bio-accumulate pesticides, such as 1,1'-(2,2,2-trichloro-1,1- ethanediyl)bis(4-chlorobenzene) (4,4’-DDT) directly through the water. Bio- concentration within fish may take place through the food chain, since many species have both plant and animal based diets, resulting in the addition of 4.4’-DDT from various levels within the same food chain. The manufacture and large scale use of 4,4’-DDT began in 1939 and ended when it was banned by the United States Environmental Protection Agency (USEPA) in 1972. This 33 year time period resulted in massive inputs of 4,4’-DDT into the natural environment. Concentrations of 4,4’-DDT are still present in areas including Europe and the United States of America where 4,4’-DDT is no longer applied in any way, shape, or form.

The current use of 4,4’-DDT in approximately 43 African countries only occurs under special dispensation from the World Health Organisation (WHO) for the control of the Anopheles sp. mosquito. This genus of mosquito is the vector for malaria, a parasitic blood infection which caused the death of 445 000 people in Africa in 2016. Current application of 4,4’-DDT in areas most affected by malaria takes place through indoor residual spraying (IRS) of houses and homesteads. The potential for 4,4’-DDT to enter surrounding aquatic systems through spray drift during IRS or runoff caused by rain following IRS is extremely high. The lipophilic nature of 4,4’-DDT and the fact that it has been established as a known endocrine disrupting chemical (EDC), makes understanding the immediate effects that 4,4’-DDT may have on aquatic biota in areas where its use is still common practice, very important.

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The focus fish species selected for this study was Synodontis zambezensis (S. zambezensis). This species was chosen due to its importance as a source of protein to communities surrounding major river systems throughout Africa and within South Africa. The higher than normal fat content of S. zambezensis means that they have the potential to bio-accumulate 4,4’-DDT and its degradation products to concentrations far higher than those found in other fish species. Results from a field study conducted in the lower Phongolo River and floodplain supported this assumption and the need for baseline laboratory based data arose. The aim of this section of the thesis was to establish a direct correlation between exposure concentrations of 4,4’-DDT and resulting bio-accumulation data. The concentrations used for the exposure of adult S. zambezensis were sub-lethal concentrations and ranged from 1.974 ng/mL to 63.168 ng/mL. Extraction of 4,4-DDT and its degradation products from the muscle tissue of S. zambezensis and the exposure tank water followed standardized Soxhlet and liquid/liquid extraction techniques respectively. The analysis of the muscle tissue of exposed S. zambezensis yielded interesting correlations between exposure concentrations and bio-accumulated concentrations. Bio-accumulated concentrations increased as the exposure concentrations increased, but the percentage of 4,4’-DDT equivalents accumulated within the fish tissue decreased. This may have been due to the increased mucal secretions present in the tanks containing the higher concentrations of 4,4’-DDT. It was clear from the results of this section of the study that bio-accumulation occurs in the muscle tissue of S. zambezensis after an acute exposure directly from the aquatic environment with no input from the food chain. Recommendations for future studies include the completion of additional acute exposures on adult S. zambezensis as well as the analysis of any resulting mucal secretions within the exposure tank water.

Biomarkers of effect and exposure indicate the physiological effects and adaptations that take place in organisms in response to toxicants. Biomarkers of oxidative stress selected for this study included superoxide dismutase (SOD), catalase (CAT), reduced glutathione (GSH), protein carbonyls and malondialdehyde (MDA). These biomarkers are a subsection of the biomarkers of effect and give an indication of changes occurring within organisms in response to the effects and by-products of reactive oxygen species (ROSs). The resulting oxidative stress that occurred in both

xix | P a g e the gill and liver tissue of S. zambezensis was a clear indication of the effects of 4,4’- DDT. The significantly higher (p≤0.05) GSH concentration found in the gill tissue at the highest 4,4’-DDT exposure concentration is a clear indication of oxidative stress as a result of exposure to 4,4’-DDT. Bimodal responses seen in the other oxidative stress biomarkers indicated that although the effects of 4,4’-DDT are clear, the fact that the exposure period was acute may have affected the physiological responsiveness of S. zambezensis. Oxidative stress biomarkers were not the only biomarkers included in this study. Biomarkers of exposure including acetylcholinesterase (AChE) and cytochrome P450 (CYP450) were analysed alongside the biomarkers of effect and indicated the same bimodal trend. The cellular energy allocation (CEA) giving an indication of the changes in energetics, showed no real overall differences between higher or lower 4,4’-DDT exposure concentrations. The addition of chronic exposure studies is recommended for studies in the future, alongside an additional acute exposure using the same concentrations of 4,4’-DDT investigated here.

The use of Ribonucleic acid (RNA) as a biomarker was also investigated in this study, since the effects of DDT may cause damage to the expression of specific sex linked proteins. The changes in gene expression linked with proteins can be used as an additional biomarker of effect. Techniques for the isolation of RNA were completed following standardised TRIzol Reagent and Nucleospin®RNA Isolation kit protocols. Following RNA isolation, conversion of RNA to complimentary deoxyribonucleic acid (cDNA) was completed following standardised protocols listed by ReverTra Ace® qPCR RT Master Mix with gDNA remover 1204. Primer design and application was completed following standardised techniques. The identification of a housekeeping (HK) gene for expression comparisons to be made was successfully completed. The HK gene identified was beta-actin (β-actin) and the primers established for this gene were efficient. Two other genes including transferrin and oestrogen receptor (ER) were identified, but the primers designed were unfortunately inefficient. Further extraction and analysis is recommended for a full picture of the effects that 4,4’-DDT may have at a genetic level.

The use of S. zambezensis as a biological indicator species was briefly investigated during a Water Research Commission (WRC) project. This study aimed to expand on the knowledge base of this species as an alternative biological indicator species.

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Two other species, namely Oreochromis mossambicus (O. mossambicus) and Poecilia reticulate (P. reticulata), which are commonly used in toxicological investigations, were selected for the determination of species sensitivity distributions

(SSDs) of the juvenile (14-day old) fish. The comparisons of calculated LC50 concentrations were completed with the use of Probability unit (Probit) analysis and were aimed at determining how sensitive S. zambezensis is compared to O. mossambicus and P. reticulata. Comparisons were also made between the current available LC50 values of Oncorhynchus mykiss (Rainbow trout), Pimephales promelas (Fathead minnow) and Lepomis macrochirus (Bluegill) which range from 3.4 ng/mL to 10 ng/mL. It was concluded that S. zambezensis is much more sensitive than both O. mossambicus, P. reticulata and the species listed previously in the Material Safety Data Sheets for 4,4’-DDT, with an LC50 concentration of 1.94 ng/mL. Recommendations for future studies include the completion of additional LC50 experiments and the completion of a Fish Embryo Test (FET).

This study confirms that S. zambezensis can be successfully used as an alternative bio-indicator species in both field and laboratory studies. This species occurs in many river systems and their use as a laboratory indicator species makes these studies easy and efficient to complete. This species is so much more sensitive during its juvenile life stages than during its adult life stages that potential effects of xenobiotics on earlier life stages could be more detrimental than previously assumed. Techniques of artificially breeding S. zambezensis have been successfully established and this will allow for further analyses of both juveniles and embryos. The aim of increasing the knowledge base surrounding this fish species has been achieved through the completion of this study, but there are many more studies, including completion of RNA analysis, increased exposure experimentation and growth and development studies, that have the potential to follow on from this study. Improved analysis of a variety of other biomarkers of both the blood and the reproductive organs should be completed to allow a more rounded understanding of the physiological functioning of this species.

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Chapter 1: General introduction

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1. General introduction 1.1 Background The introduction of pesticides into aquatic ecosystems is of great concern and leads to contamination of freshwater systems and the organisms that inhabit them (Ansara-Ross et al., 2012). Many organochlorine pesticides (OCPs) are persistent in the environment. Use of many OCPs, which began in the early 1900s, and their characteristic inability to be broken down, have led to their accumulation in aquatic ecosystems (Ansara-Ross et al., 2012). Some toxicants specifically those that are lipophilic e.g. 1,1'-(2,2,2-trichloro-1,1-ethanediyl)bis(4-chlorobenzene) (DDT), have the ability to avoid excretion from the body of the organism, by dissolving in lipid droplets found within muscle tissue and other organs (Vernouillet et al., 2010). This ability makes these substances persistent within aquatic organisms and are therefore known as persistent organic pollutants (POPs). 4,4’-DDT falls into this group and it has the potential to bio-accumulate within biota, potentially exceeding the limits at which they can be managed by homeostatic mechanisms (Ruus et al., 2005).

4,4’-DDT has been a popular topic of conversation in scientific circles since its banning in South Africa (Deribe et al., 2011) in 1996, and is well known for its use in malaria vector control programmes in many African countries. Detrimental effects on biota have been shown within many wild bird species in the form of eggshell thinning (Tucker and Haegele, 1970). The lack of calcium deposition in eggshells has led to a decrease in viable eggs and a decline in overall populations of bird species including Caracara sp. and avian fish predators (Hickey and Anderson, 1968). Some work that has been done in South Africa on Oreochromis mossambicus (O. mossambicus) and Clarias gariepinus (C. gariepinus) include laboratory exposures to DDT as well as field data collected from Roodeplaat Dam in the Gauteng province (Marchand et al., 2012). Although laboratory exposure was done, there was no biomarker analysis applied to these samples and only histological alterations in gonads and effects on sperm motility and breeding success were recorded (Mlambo et al., 2009). The present study will therefore be the first data set collected from laboratory based exposures using 4,4’-DDT as the exposure chemical, showing resulting physiological responses through biomarker analysis in Synodontis zambezensis (S. zambezensis).

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1.2 Biological markers and selected indicator species Biological markers or biomarkers, are often used during assessments of aquatic systems to determine the current physiological state of the organisms living in these systems (Carignan and Villard, 2002). Physiological responses that occur normally within organisms as an adaptive mechanisms to changes caused by toxicant exposure, have been studied at length throughout many different aquatic systems (Cormier and Daniel, 1994). The responses documented within target organisms such as fish and invertebrates, have been noted as the body’s way of dealing with or adjusting to changes within the environment (Van der Oost et al., 2003). Specific compounds which are involved in normal functioning of organisms are excited or inhibited in response to xenobiotics or their effects and these changes are used as biomarkers within organisms (Martin and Black, 1998). Biomarkers are defined as cellular or biochemical changes that are measurable in a biological system or sample. Biological response which take place in response to pollutants or stressors that can be measured at the sub-individual level and indicate a deviation from the normal state, which cannot be detected in the intact organism, is known as a biomarker (Van der Oost et al., 2003).

The species chosen for this study is the Brown squeaker, Synodontis zambezensis (Peters, 1852). This fish belongs to the family Mochokidae, the largest catfish family in Africa with the greatest distribution throughout the African continent. The reason for choosing S. zambezensis for this study is because it has been noted, through social studies, as an important source of food for toddlers and babies in the communities surrounding the lower Phongolo River and floodplain (Coetzee et al., 2015). The lack of a bony skeleton, besides those bones found within the head plate, makes S. zambezensis easy to consume for younger individuals (Smit et al., 2016). Understanding the fate of the 4,4’-DDT entering the system is extremely important and the effects that 4,4’-DDT has on these organisms will determine the ability to maintain a sustainable population in the wild for communities that rely on this species.

Despite the importance of this species as a food source, there is a distinct lack of published information on biomarker and toxicological analysis of S. zambezensis. A large majority of the published data for this genus focuses on their feeding habits (Akombo et al., 2014), taxonomy (Willoughby, 1994), morphology (Bruwer and Van

3 | P a g e der Bank, 2003), biology (Lalèyè et al., 2006) and artificial breeding (Kaiser and Rouhani, 1999; Van der Waal, 1986). A recent Water Research Commission Report (WRC Report Number: 2185/1/16) reported on bio-accumulated pesticides, including 4,4’-DDT and its degradation products as well as resulting biomarker responses in S. zambezensis sampled in the Phongolo River and floodplain (Smit et al., 2016). This addressed the question of the ability of S. zambezensis to accumulate high concentrations of 4,4’-DDT and its degradation products, but since the report was done in the field with no determination of the length of time of exposure, the resulting bioconcentration potential of the species was unclear. Biomarker analysis resulting from the above mentioned WRC study also allowed a better understanding of the physiological adaptability of S. zambezensis, but since the sampling was again field based, establishing whether the physiological adaptations indicated through the biomarker responses are inked directly to the toxic effects of 4,4’-DDT is unclear. The study by Smit et al. (2016) revealed that S. zambezensis can be used as an indicator species instead of using more well-known species such as C. gariepinus and O. mossambicus.

1.3 Motivation for study Baseline readings of the physiological responses experienced by S. zambezensis at specific concentrations of 4,4’-DDT will help in establishing the toxicity experienced by this fish species after an acute exposure under laboratory conditions.

The results will indicate, through physiological responses, the effects of 4,4’-DDT on the biological functioning of the S. zambezensis as seen in previous studies (Van der Oost et al., 2003, 1996). Once this baseline has been established, new measures may be implemented in areas where 4,4’-DDT is sprayed in the control of Anopheles sp. mosquitos. The measures put in place may lead to a decrease in the concentrations used for malaria control, thereby decreasing the effects that 4,4’-DDT may have on wild fish populations (Biggar et al., 1967; Bouwman et al., 2011). Decreases in the concentrations of 4,4’-DDT may be monitored to ensure that they are less harmful to the environment, but still effective in the control of Anopheles sp. mosquitos. The decreased concentrations of 4,4’-DDT entering the environment may also lead to a decrease in the concentrations taken in by communities using fish as a source of protein (Tomza-Marciniak and Witczak, 2009).

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A laboratory based assessment of bio-accumulated concentrations of 4,4’-DDT, biomarker analysis and LC50 determination has never been investigated in S. zambezensis. This laboratory based study will provide new information about environmentally relevant concentrations of 4,4’-DDT and the responses they elicit in fish (Smit et al., 2016). It is expected that the results from this project will enable a better understanding of the toxic effects of 4,4’-DDT at various concentrations. These effects will be reflected in changes seen within the biomarker responses (Van der Oost et al., 2003). The results will improve the efficiency of management plans for areas where spraying of 4,4’-DDT takes place and will allow for dosages to follow specific guidelines to reduce toxicity to aquatic organisms and communities in affected areas (Bouwman et al., 2011), while retaining their effectiveness against Anopheles sp. mosquitos.

The increased excitation of oxidative stress biomarkers such as malondialdehyde (MDA), superoxide dismutase (SOD), catalase (CAT), reduced glutathione (GSH) and protein carbonyls (PC) is expected as the exposure concentrations of 4,4’-DDT increase. An increased inhibition and excitation of acetylcholinesterase (AChE) and cytochrome P450 (CYP450) respectively, is expected as the exposure concentrations of 4,4’-DDT increase (Levine et al., 1990; Mateos et al., 2005; Slaninova et al., 2009; Bucheli and Fent, 1995; Leibel, 1988). Changes in energetics including an increase in energy consumption (Ec) and decreased energy availability (Ea) leading to overall decreases in the cellular energy allocation (CEA) are expected as the exposure concentrations of 4,4-DDT increase (De Coen and Janssen, 2003; Verslycke et al., 2004). Ribonucleic acid (RNA) will be used as an added biomarker of effect, and changes within the expression of focus genes are expected to be seen throughout the exposure concentrations. Higher concentrations of 4,4’-DDT are expected to elicit more pronounced inhibition or excitation of enzymes within the genes of S. zambezensis (Tom and Auslander, 2005).

1.4 Hypotheses, aims and objectives Synodontis zambezensis was chosen as the bio-indicator species for this study because of its previous successful use as a bio-indicator species (Smit et al., 2016). The ecological importance of this species as well as its importance in communities as a source of food throughout Africa (Baras and Laèyè, 2003; Coetzee et al., 2015; Daget et al., 1991), further motivates its use as a bio-indicator species. The link

5 | P a g e between communities that rely on fish as a source of food, and the effect that eating pesticide laden fish may have on communities, needs to be better understood. This study hopes to add to the understanding of the link between these two important environments.

1.4.1 Hypotheses 1. The exposure of adult S. zambezensis to varied concentrations of 4,4’-DDT in a controlled laboratory environment will lead to alterations in biochemical responses with increased excitation or inhibition seen in the higher exposure concentrations. 2. Ribonucleic acid (RNA) expression (increase or decrease) following the exposure of S. zambezensis to varied concentrations of 4,4’-DDT, will be proportional to the exposure concentrations used. 3. The exposure of adult S. zambezensis to varied concentrations of 4,4’-DDT in a controlled laboratory environment will lead to accumulation within the tissue of fish and these accumulated values will correlate with the exposure concentrations used during the 96-hour exposure. 4. The exposure of 14-day old S. zambezensis to varied concentrations of 4,4’-DDT

in a controlled laboratory environment will allow for the calculation of LC50 values for this species and higher percentage mortality will occur at higher exposure concentrations. 5. The exposure of 14-day old S. zambezensis, O. mossambicus and Poecilia reticulata (P. reticulata) to varied concentrations of 4,4’-DDT will allow for the

calculation of the LC50 values for all three fish species and subsequent species sensitivity distributions (SSDs) for each of the three species.

1.4.2 Aims This study had the following aims:

1. To determine whether exposure to 4,4’-DDT during an acute exposure period of 96-hours will elicit any notable biochemical or genetic changes. 2. To determine the bio-accumulation of 4,4’-DDT in S. zambezensis through organic extraction of 4,4’-DDT from fish muscle tissue.

3. To determine the lethal concentration at which 50% mortality occurs (LC50) for 4,4’-DDT in S. zambezensis, O. mossambicus and P. reticulata.

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4. To determine the species sensitivity distribution (SSD) for each of the three species compared to previous studied species as well as the three species used during this section of the study.

1.4.3 Objectives The aims of the study are listed above and each of these aims were achieved through the protocols and procedures set out in the following objectives:

i. Exposure of adult S. zambezensis and juvenile S. zambezensis, O. mossambicus and P. reticulata to varied concentrations of 4,4’-DDT. ii. Analyses of bio-accumulation of 4,4’-DDT in fish tissue and exposure tank water. iii. The use of biomarkers of exposure to determine the extent of 4,4’-DDT (AChE and CYP450) exposure in S. zambezensis. iv. The use of biomarkers of effect to determine changes in cellular energy (CEA) within muscle tissue and oxidative stress (SOD, PC, MDA and CAT) in liver and gill tissue of S. zambezensis. v. Analysis of toxico-genomic changes with the use of RNA of a biomarker in S. zambezensis. vi. Application of multivariate statistical analyses (Quinn and Keough, 2002) to determine whether links are present between 4,4’-DDT bio-accumulation and biomarker responses in S. zambezensis. vii. Application of transformative Probability Unit (Probit) analysis (Finney, 1947;

Finney and Stevens, 1948) to percentage mortality data collected during the LC50

exposure experiments for the determination of LC50 concentrations in the three focus fish species of this section of the study.

1.5 Layout of thesis This thesis is split into the following chapters. Each chapter will outline a specific topic focussed on during the study and will contain its own introduction, materials and methods and results and discussion sections. The references for each of these chapters will be combined and included into one reference chapter at the end of the thesis.

Chapter 1: General introduction This chapter aims to introduce the overall topic to the reader and allow for brief introduction to the topics and themes that will be elaborated on during the chapters

7 | P a g e to follow. The motivation for the study as a whole has been discussed in this chapter with focus on the use of S. zambezensis as a bio-indicator species as well as some previous research on the species. The gap that will be filled by this thesis has been explained during this chapter and the hypotheses, aims and objectives for the study are listed.

Chapter 2: Literature review The focus chemical, 4,4’-DDT is discussed fully in this chapter. The sections relating to 4,4’-DDT include its historical and current use in South Africa and abroad, its intended target and the fate of 4,4’-DDT in the environment. The bio-accumulation, regulation and absorption sections of this chapter clearly outline the processes that lead to the effects of 4,4’-DDT on fish and their offspring. The framework, in which all of the toxicity testing for this study will be completed, is discussed in full. Biomarkers of exposure and effect as well as RNA as a biomarker are outlined and explained not only in terms of what they indicated but also in terms of previous work that has been successfully completed using these biomarkers. A short introduction to each of the sections of the thesis is included in this chapter, as well as a literature review of the chapters and topics to follow.

Chapter 3: Focus species The species chosen for this study S. zambezensis, is discussed in detail in this chapter, with details on their habitat preferences, distribution and feeding habits being included. The importance of this species as a source of food throughout Africa is discussed here as well as its taxonomy, morphology and importance as a possible laboratory test organism.

Chapter 4: Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852) The article published in Aquaculture Research, “Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852)”. This chapter outlines the artificial breeding procedures followed, housing and care of S. zambezensis. The juveniles bred using these procedures were used in the LC50 exposure experiments to follow.

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Chapter 5: Baseline bio-accumulation concentrations and resulting oxidative stress in Synodontis zambezensis after an acute laboratory exposure to 4, 4’- DDT This chapter was published online in Pesticide Biochemistry and Physiology on the 6th of February 2019. This chapter includes the resulting bio-accumulation data as well as the noted oxidative stress (SOD, MDA, CAT, PC and GSH) responses found within the gills and liver tissue of S. zambezensis.

Chapter 6: Biomarkers of exposure and energetics This chapter contains the remaining biomarker data. The biomarkers of exposure, acetylcholinesterase (AChE) and cytochrome P450 (CYP450) are discussed in relation to the bio-accumulation data listed in detail in chapter 5.

Chapter 7: Ribonucleic acid (RNA) as a biomarker The techniques used and completed for the analysis of genetic changes in S. zambezensis as a result of exposure to 4,4’-DDT are discussed fully in this chapter. Isolation of RNA, conversion to DNA (cDNA), the procedures and techniques used to design successful primers are all discussed in this chapter. The use of RNA as a biomarker of toxicant induced genetic change is the main focus of this chapter.

Chapter 8: LC50 exposures and species sensitivity distributions The resulting LC50 values and species sensitivity distributions (SSDs) for S. zambezensis, O. mossambicus and P. reticulata are discussed in this chapter. The application of transformative Probit analysis and its history and importance in the toxicological field of study, as well as equations and statistical analysis required for the determination of LC50’s are the focus of this chapter.

Chapter 9: Conclusions and recommendations The final conclusions and recommendations are listed in this chapter. Possible investigation that may build on the data gained during this study as well as recommendations for improvement on experimentation completed during this study.

Chapter 10: References The references for the entire thesis will be listed in this chapter to avoid repetition of references throughout the thesis.

Appendix A: Raw data This section contains the raw data of the bio-accumulation and biomarkers analysed during this study.

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Appendix B: Published papers The front page of published papers from this thesis are included in this section.

Appendix C: Conference contributions All contributions to both national and international conferences are listed in this section.

1.6 Anticipated contribution This project will contribute to filling the toxicological knowledge gap that currently exists for S. zambezensis. It will allow for direct linkages to be made between cause and effect. The relationship between the concentrations of 4,4’-DDT used during the exposures and the resulting physiological adaptation seen in the biomarkers of exposure and effect will be clearly illustrated by this study. These linkages will allow for better interpretation of biomarker responses resulting in tissue collected during future field based studies. The LC50 values of three fish species will add to the toxicological data for 4,4’-DDT and increase the material safety data sheet (MSDS) information. The addition of these three species to the MSDS information may allow for a more holistic understanding of the toxicological effect of 4,4’-DDT on fish species that occur in different trophic niches within natural environments. Further species sensitivity distribution studies can be done including these three species of fish.

The importance of S. zambezensis as a biological indicator species is crucial since this species is of economic importance throughout Africa. A manuscript entitled “Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852)” has already been published in Aquaculture Research and this manuscript will contribute to the knowledge on artificial breeding of this species and other species within the genus. The maintenance of this species in a laboratory environment and subsequent use as a biological indicator species will allow for important links to be made between human health and toxicant exposure.

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Chapter 2: Literature review

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2. Introduction 2.1 4,4’-DDT 1,1'-(2,2,2-trichloro-1,1-ethanediyl)bis(4-chlorobenzene) (IUPAC) more commonly known as 4,4’-DDT is a crystalline insecticide (WHO, 1979). Its physical properties include being colourless, tasteless, odourless and extremely lipophilic. The preparation of 4,4’-DDT is a relatively inexpensive and simple process, although it requires being dissolved in an organic solvent due to its hydrophobic nature (Mellanby, 1992). 4,4’-DDT is synthesized through hydroxylation reactions between chloral (CCl3CHO) and chlorobenzene (C6H5Cl) in the presence of an acidic catalyst

(H2SO4). Technical grade 4,4’-DDT, used for malaria vector control of the Anopheles sp. mosquito, is composed of almost 14 chemical compounds (Metcalf, 1995). The active ingredient 4,4’-DDT makes up 65-80%, with 2,4’-DDT making up 15-21%. The remaining components include 4% 4,4’-DDD and 1.5% 1-(p-chlorophenyl)-2,2,2- trichloroethanol (Metcalf, 1995). 4,4’-DDT has a total of six metabolites which are divided between two isomers. The main isomer is the 4,4’-isomer and the other is the 2,4’-isomer (WHO, 1979). The metabolites are therefore DDT, DDE and DDD (Figure 2.1) in each of the isometric forms. The compound 4,4’-DDT (98.7%, Sigma- Aldrich) was used for the exposures in this study.

Figure 2. 1: Degradation of 4,4’-DDT to form its metabolites, 4,4’-DDE (removal of HCl) and 4,4’-DDD (through reductive dechlorination) (Li et al., 2010).

2.1.1 Historical use of 4,4’-DDT The initial synthesis of 4,4’-DDT took place in 1874 by the Austrian chemist, Othmar Zeidler (Zeidler, 1874), but its properties as an insecticide were only discovered in 1929 (Brand et al., 1930; Brand and Bausch, 1930). Paul Hermann Müller, a Swiss

12 | P a g e chemist discovered its effectiveness and was awarded the Nobel Prize in Physiology or Medicine in 1948 for “his discovery of the high efficiency of 4,4’-DDT as a contact poison against several arthropods" (Nobelprize.org, 1948).

Large scale use of 4,4’-DDT was first implemented during World War II (WWII) for the control of fleas and lice, which spread typhus, in both troops and civilians. Images from that time period show children running behind trucks travelling down roads in villages, spraying fine mists of 4,4’-DDT throughout affected neighbourhoods (Dunlap, 1981). After the war, 4,4’-DDT became available to civilian health agencies largely because 80% of all infectious diseases that affect human populations are spread by insect vectors or other small arthropods (Soulsby and Harvey, 1972). The effectiveness of 4,4’-DDT led to the almost complete eradication of typhus in Europe (Gladwell, 2001). Aerial spraying of 4,4’-DDT was introduced to control the spread of mosquito borne diseases such as malaria and dengue fever, and as a result, Europe and North America are malaria free to this day. The use of 4,4’-DDT in many households was common practice for the control of pests, and the application to crops such as corn and maize was the main defence against produce destroying insects (Davies et al., 2007).

The World Health Organization (WHO) implemented a large scale operation with the aim of complete eradication of malaria in countries with moderate to low transmission rates; 4,4’-DDT was their main tool in this operation. The success of this programme meant the total extermination of the disease in Taiwan, a large portion of the Caribbean, some parts of northern Africa and Australia and the Balkans (De Zulueta, 1998). The efficiency of 4,4’-DDT was further noted with the complete annihilation of malaria in a large part of the South Pacific and a significant decrease in mortality in Sri Lanka and India. The initial success of this operation was sustained in areas with high socio-economic status, but malaria has since resurfaced in poorer areas, and because mosquitos have developed a tolerance to 4,4’-DDT, some areas have experienced an increase in transmission and infection (Sadasivaiah et al., 2007).

The publishing of the book Silent Spring by Rachel Carson in September 1962 eventually led to the banning of 4,4’-DDT in America by 1972 (Lear, 1998). The ban was instituted by the United States Environmental Protection Agency (USEPA) after

13 | P a g e investigation into the claims of the “toxic and carcinogenic characteristics” of 4,4’- DDT and the threat to animal and human life alike (Attaran et al., 2000).

2.1.2 The use of 4,4’-DDT in South Africa Well known for its use in controlling the spread of malaria in parts of Sub-Saharan Africa, 4,4’-DDT has also been used since the 1970s in the agricultural industry until it was outlawed for this specific purpose in 1974 (Bouwman, 2004; Van Dyk et al., 2010). In spite of 4,4’-DDTs outlawed status, large stockpiles of the organochlorine were still found as late as 1995, and it was commonly accepted that these stockpiles were used informally, until all use of 4,4’-DDT was completely banned by the Stockholm Convention in 1996 (Bouwman, 2004).

Malaria control programmes have effectively been using 4,4’-DDT since the 1940s. The success of these programmes can be attributed to the cheap, long lasting presence and easy application of 4,4’-DDT through indoor residual spraying (IRS) (WHO, 2002). Malaria control programmes have been credited with the complete purge of the Anopheles phenestus the main malaria vector during the 1940s and 50s (Coetzee et al., 2013). After the banning of 4,4’-DDT in 1996, pyrethroid insecticides became the replacement pesticide of choice and they proved effective against the Anopheles arabiensis species. Pyrethroid insecticides became ineffective at curbing the spread of malaria when Anopheles sp. mosquitos developed resistance to their properties, and this meant that 4,4’-DDT had to be reintroduced as the main combat to the spread of malaria (Davies et al., 2007).

The reintroduction of 4,4’-DDT into the Malaria Control Programme in South Africa occurred in 2000, when a rapid increase in infection rate occurred in KwaZulu Natal. The use of 4,4’-DDT in IRS was extended to the Limpopo and Mpumalanga provinces, even though pyrethroid insecticides have proven effective against Anopheles arabiensis, the main vector of malaria in these two provinces (Bouwman et al., 2011). The exemption lodged in 2004 with the Secretariat of the Stockholm Convention clearly states that 4,4’-DDT may not be used in areas where “locally safe, effective and affordable alternatives are available.” Under this exemption, 4,4’- DDT is only to be imported into South Africa when it is bought by the Department of Health (DOH) with clear applications and uses stated with every import.

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A health risk assessment conducted in 2006, showed the presence of 4,4’-DDT and pyretheroid residues in the breastmilk of woman living in a malaria-endemic area in KwaZulu Natal, South Africa (Bouwman et al., 2006). Residues of 4,4’-DDT were also found within the water and sediment of the Rietvlei Nature Reserve in the East of Pretoria, South Africa (Bouwman et al., 2006).

2.1.3 Intended target of 4,4’-DDT The intended target of 4,4’-DDT and other insecticides is the voltage-gated sodium channels found within the insect nerve cell membranes (Davies et al., 2007). These specialised channels are not unique to insects and are found within almost all organisms with a highly developed nervous system, including human beings (Moran et al., 2015). These specialised units are responsible for the initiation and spread of action potentials through excitable cells found in nerves, muscle tissue and neuroendocrine cells (Hille, 2001; Hodgkin and Huxley, 1952). Sodium channels consist of α units and β units (Figure 2.2), which interact with one another to allow for the transfer of ions into and out of the cell through this section of the cell membrane (Brackenbury and Isom, 2011).

Figure 2. 2: Graphical representation of the primary structure of voltage-gated sodium channel with α and β units (Catterall, 2000).

4,4’-DDT affects the ‘para’ voltage-gated sodium channel in insects and its name is derived from the paralysis locus in which it is found (Loughney et al., 1989). This insect sodium channel is structurally homologous to the α-subunit of mammalian voltage-gated sodium channels (Catterall, 2000). The peripheral nervous system is primarily affected by 4,4’-DDT and initial contact with the insecticide causes

15 | P a g e spontaneous firing of neurons which results in muscle twitches leading to tremors throughout the body. The initial exposure does not kill the insect immediately, but after a few hours or days of continuous exposure, death of the insect results from excitatory paralysis (Perry et al., 2011). These effects are caused because 4,4’-DDT is a channel agonist, which binds to open channels and does not allow them to close (Elliott et al., 1987; Ottea et al., 1990).

The overall insecticidal effect of 4,4’-DDT and its metabolites is influenced by its molecular shape and size, and for this reason only the 4,4’-isomer has significant insecticidal activity (Hutson and Roberts, 1985).

2.1.4 The fate of 4,4’-DDT in the environment The introduction of 4,4’-DDT into the environment occurs through various pathways. These include spray drift as a result of IRS (Bouwman et al., 2006) and runoff due to heavy rains after unlawful application of 4,4’-DDT to crops (Van Dyk et al., 2010). Since technical grade 4,4’-DDT is a characteristically unstable chemical (ATSDR, 2002), immediate breakdown into its metabolites occurs after contact with sunlight, moisture and increased temperatures (ATSDR, 2002). Sediments and soils act as sinks for many organochlorine pesticides (OCPs) including 4,4’-DDT, and this means that when these sediments or soils come in contact with any volatizing agents, 4,4’- DDT may enter its immediate environment from these sinks, whether this is in an aquatic or terrestrial environment. Soils and sediments are dangerous long term sources of exposure for animals (USEPA, 1979).

Aquatic ecosystems have a unique interaction with 4,4’-DDT because of its hydrophobic nature (Kidd et al., 2001). The large majority of 4,4’-DDT and its metabolites entering bodies of water are not biologically available for long periods of time. This is because as soon as 4,4’-DDT enters a water system it adheres to, is absorbed by, or adsorbs to plants, animals, sediment and suspended organic matter. This ability makes 4,4’-DDT and its metabolites even more resilient within a system, as it becomes part of the food web almost immediately in one form or another (USEPA, 1979). A microcosm study conducted in 1979 by the USEPA illustrates the fate of 4,4’-DDT in the aquatic environment. This study involved the application of 4,4’-DDT to a pond, and subsequent analysis of the dosed water up to 30 days after dosing and then again after 40 days. The results of this study clearly show that the

16 | P a g e initial 4,4’-DDT concentration had declined below detectable limits. 4,4’-DDT was present in the water mostly as 4,4’-DDT during the first 30 days, as 4,4’-DDT or DDD during the next 30 days and as DDE thereafter (USEPA, 1979). A study conducted on a freshwater lake showed that 4,4’-DDT accumulated to a higher concentration in fattier fish and fish which occupied the pelagic rather than the benthic food web (Kidd et al., 2001).

The manifestation of 4,4’-DDT in areas like the Arctic and other colder regions of the world, where it is not used at all in any shape or form, is due to a phenomenon known as global distillation (ATSDR, 2002). Also known as the grasshopper-effect, global distillation (Figure 2.3) is a process that many chemicals undergo, especially persistent organic pesticides (POPs) (Wania and MacKay, 1996). Chemicals are transported from warmer regions of the earth to colder regions of the Earth through this geochemical process (Wania and MacKay, 1996). Much the same as the cycle of water on earth, the chemicals are vaporized in the warmer regions, they then travel long distances in this vapour form until reaching colder regions where condensation takes place, releasing the chemicals once again and return them back to the Earth’s soil or water (Manchester-Neesvig and Andren, 1989). The overall outcome of this process is the net movement of chemicals from low to high altitudes or latitudes; depositing chemicals in regions which do not use them (Cousins and Jones, 1998).

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Figure 2. 3: Illustration of the “Grasshopper Effect” also known as global distillation. Adapted from Bhardwaj et al.(2018).

The formation of the DDE metabolite takes place through photochemical reactions which requires the presence of sunlight (Leaños-Castañeda et al., 2007; Maugh, 1973) as well as through bacterial dehydrochlorination (Leaños-Castañeda et al., 2007; Pfaender and Alexander, 1972) and animal dehydrochlorination (Kurihara et al., 1988). DDD on the other hand forms through a process of reductive dehydrochlorination which can be a result of chemical reactions (Baxter, 1990; Castro, 1964; Glass, 1972; Miskus et al., 1965; Zoro et al., 1974) or microbial mediation (Wedemeyer, 1966; Yao et al., 2007).

2.2 Bio-accumulation Aquatic ecosystems are polluted relatively easily, since agriculture and industry are more often than not situated very close to their water sources (Moore et al., 2002). When previously established baseline concentrations of chemicals or substances are exceeded and cause distinct negative impacts on the environment, this is defined as pollution (Koç, 2010). Establishing baseline environmental concentrations of chemicals is important, because deviations in quality and status of aquatic systems can only be detected if the baseline is already known (Roux et al., 1994).

The potential for a chemical to bio-accumulate can be predicted from the physical- chemical data of that chemical. If the log octanol-water partition coefficient (Kow) value is low, the bio-concentration factor (BCF) will also usually be low and therefore

18 | P a g e the potential to bio-accumulate would be slight. The converse is applicable when the log Kow is high (OECD, 2014). Interactions within the environment may affect the toxicity and accumulation potential of pollutants, therefore a pollutant may be more toxic in one system than in another, purely based on the physical parameters of the aquatic system (Wang et al., 2012). The physical parameters with the highest impact on pollutant toxicity are pH and temperature because they both affect the strength of the bonds present within chemicals (Giesy and Hoke, 1989). An increase in either pH, temperature or both, will increase the bio-availability of the pollutant because speciation of the pollutant increases and so too does solubility (Giesy and Hoke, 1989).

The bio-availability of a chemical is defined by Jezierska and Witeska, (2006) as the concentration of a pollutant available to biota within a system that can be absorbed. Bio-accumulation of a pollutant within an organism is determined by the nature of the pollutant (lipophilic/hydrophilic), as well as the ability of the organism to excrete the pollutant (Hosseini Alhashemi et al., 2012). Bio-accumulation can occur through bioconcentration or through bio-magnification (Bervoets and Blust, 2003; Wepener et al., 2011). Time is the determining factor when it comes to bio-concentration as the longer an organism has been exposed to a particular chemical the higher the concentration becomes within the organism (Bervoets and Blust, 2003). Bio- concentration however, is affected by the position that the organism occupies within the food chain. The higher up the organism is in the food chain, the higher their biomagnification (Wepener et al., 2011).

2.2.1 Absorption and regulation Bio-concentration and bio-magnification are directly influenced by the pathways of accumulation within the organism (Van der Oost et al., 1996). Bio-concentration pathways involve absorption of substance across membranes and body surfaces that are in direct contact with the surrounding environment, as seen in fish and amphibians (Van der Oost et al., 1996). As soon as the toxicant has entered the organism, it needs to be excreted from the organism or sequestered throughout the organism. The process of sequestration is described by Onsanit and Wang, (2011) as storage of chemicals in areas of the body that are not directly linked to physiological functioning of the organism. Areas of sequestration include hair, nails, scales and teeth.

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Lipophilic toxicants, like 4,4’-DDT, have the ability to delay or avoid excretion from the organism completely. These toxicants dissolve in the lipid droplets of organisms the result is bio-concentration (Vernouillet et al., 2010). Homeostatic mechanisms within organisms are unable to manage the uptake of these persistent substances and the concentrations will quickly exceed the limits of the biota affected (Ruus et al., 2005).

2.3 Effects of 4,4’-DDT on fish and offspring Exposure of fish to xenobiotics in the aquatic environment is especially important because they are in constant contact with their environment and this fact means that exposure can occur through more than one pathway (Smolders et al., 2003a). Direct exposure through the skin and gills occurs in the case of bio-available xenobiotics. Hydrophobic xenobiotics, like 4,4’-DDT, can be taken up indirectly through sediments and plant matter ingested during feeding (Wepener et al., 2011). Possible exposure of fish to xenobiotics is therefore inevitable and their position in the food chain also makes them specifically important since they can be a link between the pollution of aquatic environments and human populations (Chapman and Jackson, 1996).

Many surveys have been carried out in South Africa in an effort to assess the levels of 4,4’-DDT, its metabolites and other pesticides present in various aquatic ecosystems (Barnhoorn et al., 2009; Viljoen et al., 2016). A study conducted in 2009 by Barnhoorn et al. found levels of 4,4’-DDT and its metabolites present both in the water and in the tissue of various fish species in various dams in the Limpopo Province. Levels of the DDE metabolite were the highest in this study being found at four of the eight sites sampled and in both fish species O. mossambicus and C. gariepinus.

Parvez and Raisuddin, (2005) suggest that changes seen in organisms in response to pollutants at various levels including cellular and biochemical (Figure 2.4), may potentially affect the integrity of the entire population and in turn the ecosystem. 4,4’- DDT in its various forms has been shown to cause the functional sex reversal in Japanese rice fish (Medaka) of male to female (Hano et al., 2007), reduction in fecundity and population recruitment (Patyna et al., 1999) as well as reproductive dysfunction due to direct or indirect exposure to 4,4’-DDT (Kime and Nash, 1999).

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Figure 2. 4: Trade-off between response sensitivity and ecological relevance over time at the different levels of biological organisation. Adapted from Adams, (2001).

Most studies completed on the effects of 4,4’-DDT on fish, concentrate on their endocrine disrupting qualities. Since sexual development is very sensitive and responsive to the effects of oestrogenic and androgenic compounds, the fact that 4,4’-DDT mimics the effects of oestrogenic compounds means that its effects on sexual development in fish is of great importance. Interactions with physiological systems and alterations caused during development, growth and reproduction have been found in fish exposed to 4,4’-DDT (Jobling et al., 1998).

This study focuses specifically on the physiological changes at a sub-cellular level in the form of inhibition or excitation of various biomarkers. The biomarkers interpreted were studied in exposed adult fish. Although none of the biomarkers selected show specific changes in sexual development of S. zambezensis, they allow for a better understanding of the immediate physiological changes that take place within exposed adult fish and what resulting long term exposure changes could occur.

2.3.1 Developmental toxicity Various reproductive and developmental defects have been reported in a wide variety of species ranging from snails to humans and these abnormalities have been associated with environmental exposure to toxicants, primarily oestrogen mimics (McLachlan, 2001). Since 4,4’-DDT and its isomers and metabolites mimic the effects of oestrogen within the environment it may be associated with the same abnormalities. Several reviews have been conducted on the toxic effects and health

21 | P a g e risks associated with the exposure of 4,4’-DDT in humans and wildlife (Dalvie et al., 2004; Rogan and Chen, 2005; Turusov et al., 2002).

The 2,4’-DDT isomer has been shown as a very toxic oestrogen agonist during in vitro tests with oestrogen responsive tumour cells (Klotz et al., 1996; Soto et al., 1994) as well as in a yeast oestrogen screening (YES) assay (Klotz et al., 1996). Reduced development of male alligator penises has been linked to the same isomer in contaminated Lake Apopka, Florida, USA (Guillette et al., 1994). Dose-dependent atrophy of flounder testes that were treated with 4,4’-DDT was reported by Zaroogian et al. (2001). Exposure to sub-lethal concentrations of 2,4’-DDT during crucial stages of gonadal development has also been shown to profoundly affect sexual differentiation within fish species and this can in turn affect the sex ratios of the entire population within a system (Edmunds et al., 2000). The oestrogen receptor (ER) is also affected by the 2,4’-DDE isomer and is understood to have similar agonist potency to the ER as that of 2,4’-DDT (Donohoe and Curtis, 1996).

In order for reproduction to be successful within vertebrate species, viable gametes need to be available. Disturbances of gamete development can take many forms, one of which is the introduction and subsequent absorption of toxicants from the immediate environment (Kime and Nash, 1999). Should the gametes not develop past a certain stage, poor quality gametes will result, and this will lead to low fertilisation success or a low survival rate of the offspring (Kime and Nash, 1999). Continued exposure of surviving embryos following successful fertilisation, may cause further complications in maturation of offspring. Even if the affected offspring develop to maturity, they may have a very low reproductive potential because of their exposure to pollutants during early developmental stages (Kime and Nash, 1999).

2.4 Toxicity testing framework All toxicity testing completed during this study was performed in alignment with the relevant Organisation for Economic Cooperation and Development (OECD) guidelines for testing the effects of chemicals on fish in various life stages in controlled environments (OECD, 2014).

2.4.1 Acute exposure Acute exposures are designed to take place over a period of 96 hours with mortalities of fish being recorded at 24 hour intervals. These types of exposures are

22 | P a g e used to determine the lowest concentration of a specific toxicant at which 50% of the exposed individuals die (LC50) (OECD, 2014). The test organisms should be readily available throughout the year, easy to maintain in a controlled environment, convenient to test and have biological or economical relevance (Zhou et al., 2008). Physical health of the fish should be good and they should have no visible malformations (Zhou et al., 2008).

2.4.2 Acute fish toxicity test (OECD TG, 203) Acute toxicity testing using vertebrates, in this case fish, is used for regulatory purposes of chemicals used for plant protection and pharmaceuticals (Scholz et al., 2013). Regulatory authorities often require fish toxicity data for three main reasons: 1) hazard classification of chemicals, 2) identifying whether waste products can cause fish mortalities and 3) these tests are used in conjunction with tests on other fish species as well as Daphnia sp. to give an extrapolated estimation of the chronic no-effect concentrations (OECD, 2014).

Fish testing is also employed because it is important to know whether the focus chemical has the potential to bio-accumulate within fish (OECD, 2014). The bio- accumulation potential of a chemical has the potential to cause harm to fish populations as well as human populations and other wildlife (OECD, 2014).

All of the developed OECD guidelines address one or more of the objectives in order to gain insight into the acute or long-term toxicity or bio-accumulative behaviour of chemicals in fish, with the aim of protecting the continuing sustainability of aquatic species (OECD, 1992). The data supplied by these tests allow for the completion for hazard and risk assessment of top predator species within aquatic food chains and possible secondary poisoning that could result in human populations exposed via aquatic food chains (OECD, 1992).

2.5 Biomarkers Biomarkers are used as indicators of change within an organism at a cellular level, and these changes can, more often than not, be linked to exposure to chemicals that are foreign to a specific environment (Rola et al., 2012). The changes seen can be early warning signs of further changes which may occur at higher levels of biological organisation (Wepener et al., 2005) (Figure 2.4). Changes in molecular biomarker values are early signs of change which can result after exposure, absorption and

23 | P a g e regulation of pollutants that are biologically available within aquatic ecosystems (Van der Oost et al., 1996).

Two categories of biomarkers have been selected, i.e. biomarkers of exposure as well as biomarkers of effect. The exposure biomarkers are enzymes; Cytochrome P450 (CYP450 – responds to organic compounds e.g. OCPs) (Burke and Mayer, 1974) and acetylcholine esterase (AChE – responds to pesticide exposure) (Ellman et al., 1961) that reflect signs of exposure to toxicants within the immediate environment. The effect-biomarkers principally reflect the oxidative status of cells through the use of enzymes such as catalase (CAT) (Cohen et al., 1970) and superoxide dismutase (SOD) (Greenwald, 1991). The oxidative status of cells can also be indicated by the presence of stress by-products such as malondialdehyde (MDA) (Ohkawa et al., 1979), protein carbonyls (PC) (Parvez and Raisuddin, 2005), as well as the non-enzymatic reduced glutathione (GSH) (Cohn and Lyle, 1966). The cellular energy allocation (CEA) (De Coen and Janssen, 1997) biomarker is an indication of cellular energy utilisation during stress conditions. All biomarker analysis has been completed following the standardised techniques mentioned above.

2.5.1 Biomarkers of exposure Cytochrome P450 (CYP450) Cytochrome P450 is a Phase 1 bio-transformation detoxifying enzyme and is responsible for initiation, conjugation and purge of xenobiotics from the body (Kirby et al., 2004; Schirmer et al., 2001). An assortment of xenobiotics, including polyaromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) are detoxified by this mixed-function oxygenase system (MFO) (Kirby et al., 2004). Lipophilic compounds are transformed into hydrophilic compounds through the insertion of a single oxygen atom derived from O2 (Gonzalez, 1988) allowing the substance to be excreted from the body (Van Dam et al., 1999). Cytochrome P450 forms part of the normal functioning of many organisms and plays a major role in the regulation of endogenous substances, including steroids and vitamins (Kirby et al., 2004). Since normal functioning of growth and reproduction can be adversely affected by increased levels of CYP450, it is used extensively as a biomarker and is easily induced by exposure to pollutants (Fang et al., 2010; Kosmala et al., 1998).

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Acetylcholinesterase (AChE) Found both in vertebrate and invertebrate organisms, AChE is an enzyme that is related with the neuromuscular junction (Solé et al., 2010). The role of this enzyme is in breaking down the neurotransmitter acetylcholine (ACh) which is responsible for the transfer of impulses involved in rudimentary organism functioning. These functions include swimming, feeding, respiration, hormone transfer and reproduction (Solé et al., 2010). Inhibition of AChE results in a surplus of ACh in the neuromuscular junction which in turn results in overstimulation of the synapse. This overstimulation can cause malfunctions which manifest as changes in behaviour and ultimately death of the organism (Walker, 2006).

Acetylcholine is broken down through cleavage of choline and acetate and has been used as a well-known biomarker, indicating contamination by several environmental contaminants (carbamate insecticides, methanol and heavy metals) (Rico et al., 2007; Senger et al., 2006; Yi et al., 2006). Acetylcholinesterase activity as a biomarker has also formed part of monitoring systems specifically focused on changes in aquatic organisms (De la Torre et al., 2002).

Cellular energy allocation (CEA) Energy-consumption, manipulation and storage are referred to as metabolism (Nelson, 2005). Only a certain fraction of the energy consumed in food, is absorbed by an organism across the intestinal epithelium and the remainder is lost through secretions (Nelson, 2005). The use of cellular energy allocation (CEA) as a biomarker indicating the effects that toxic stress has on the energetics of an organism has been used successfully in various studies (De Coen and Janssen, 2003). Biochemical methods are used to compare the energy consumption (Ec), and energy availability (Ea) which together gives an overall energy budget available to the organism (De Coen and Janssen, 2003).

The aim of any living organism is to maintain a stable internal environment through homeostasis (Verslycke et al., 2004). Changes to the homeostasis of an organism can occur through the metabolism of toxic substances and this can in turn affect the energy allocation (Atli et al., 2006). Re-allocation of energy from processes such as reproduction and growth to the breakdown of toxicants results in changes within the organism’s energetics (Smolders et al., 2003b). Measuring the overall energy budget

25 | P a g e can therefore serve as a sensitive biomarker of changes in energy caused by stress (Smolders et al., 2003b).

The Ea is the amount of stored energy present in an organism, which is made up of protein, glucose and lipid content. Energy consumption is calculated through measuring the amount of activity in the electron transport system (ETS) within the mitochondrial organelles of cells, in other words how much energy is being consumed through metabolism. The overall energy budget is calculated by subtracting the consumed energy from the stored energy (De Coen and Janssen, 2003).

2.5.2 Biomarkers of effect Oxidative stress biomarkers The formation of reactive oxygen species (ROSs) within organisms, occurs as a standard by by-product of cellular respiration (Ray et al., 2012), but the imbalance of these ROSs can have negative effects on organs and organ systems. Severe damage is caused to Deoxyribonucleic acid (DNA) and cell membranes by ROSs and this damage can lead to cell death and inevitably death of tissues (Wiseman and Halliwell, 1996). Within fish there are 3 specific ROSs which form in response to exposure to xenobiotics. These ROSs are hydrogen peroxide, superoxide radicals and hydroxyl radicals, all of which result in oxidative stress within fish (Atli et al., 2006). The antioxidants synthesised in order to cope with the formation of these ROSs are superoxide dismutase (SOD) and catalase (CAT). They act by removing the ROSs from cells through various processes (Atli and Canli, 2008; Prieto et al., 2007).

Superoxide dismutase (SOD) Lipid peroxidation is the breakdown of the fatty components of cell membranes and is defended against primarily through the action of the antioxidant SOD. The superoxide anion is changed into hydrogen peroxide and oxygen through the action of SOD (Marklund and Marklund, 1974). The up-regulation of SOD occurs in response to oxidative stress within oxygen utilising organisms (Choi et al., 1999) and is therefore commonly used as a biomarker of oxidative stress (Parolini et al., 2010).

Catalase (CAT) Catalase works in conjunction with SOD in that it changes the hydrogen peroxide formed during the action of SOD, into water and oxygen which are completely

26 | P a g e unreactive. The SOD – CAT pathway becomes active during aerobic metabolism or after the induction of ROSs due to exposure to pollutants (Khessiba et al., 2005). The liver and kidney peroxisomes are the only place in fish that CAT enzymes are found (Khessiba et al., 2005). Metals and pesticides may cause an increase or decrease in the concentrations of CAT present in the organism; this change is dependent on the concentrations of the pollutants present, the species affected and the route of uptake of the organism (Atli et al., 2006). Changes in pH, temperature and the presence of parasitic infections can also cause changes in CAT concentrations (Garcia et al., 2011). Catalase is generally considered to be a sensitive biomarker of environmental changes and stressors (Vasylkiv et al., 2011).

Malondialdehyde (MDA) Polyunsaturated fatty acid oxidation results in the formation of the by-product, malondialdehyde (Flohr et al., 2012). Oxidation of these fats is a result of lipid peroxidation as a result of increased stress due to ROSs (Flohr et al., 2012). Cell functioning is compromised in the presence of lipid peroxidation since the integrity of the cell membranes is affected and results in an increased membrane permeability (Flohr et al., 2012). Lipid hydroperoxide and malondialdehyde are the two main by- products formed when cell membranes begin to breakdown (Mateos et al., 2005) and for this reason MDA is an effective biomarker (Halliwell and Gutteridge, 2015; Martínez-Alfaro et al., 2006; Wu et al., 2012).

Protein carbonyls (PC) Parvez and Raisuddin, (2005) describe the structural changes in amino acid groups present on proteins. These changes occur as a result of protein peroxidation and can alter the function of the protein within the body. The attachment of carbonyl groups to the affected proteins is caused by ROSs; damaged amino acids or proteins are known as protein carbonyls (Stadtman and Berlett, 1998; Zusterzeel et al., 2001). Protein carbonyl formation is irreversible and can therefore be used as a biomarker of oxidative stress (Almroth et al., 2008, 2005).

Reduced glutathione (GSH) Mainly present in the cells in its reduced form, GSH is a tripeptide that acts through the donation of electron pairs to charged or neutral chemical species (Atli and Canli, 2008). The vast functioning of GSH in an organism ranges from the production of proteins and DNA to defence against potentially harmful xenobiotics through

27 | P a g e oxidative reactions and free radical scavenging (Peña-Llopis et al., 2001). However, the content of GSH within fish tissue may show both increases and decreases when exposed to harmful xenobiotics since they are known to have organ-specific responses to such pollutants (Ali et al., 2004; Peña et al., 2000; Sayeed et al., 2003). Levels of GSH are a useful tool when determining if damage to cells has occurred and has been successfully used as a biomarker in many previous studies (Atli and Canli, 2008).

2.5.3 RNA as a biomarker Gene products affected by xenobiotic activity are being used more often as biomarkers of environmental contamination (Tom and Auslander, 2005). Expression patterns of focus genes are quantified and the changes in expression seen are then compared with concentrations of toxicant exposure within the environment (Tom and Auslander, 2005). It is possible that all of the previously mentioned biomarkers may be transcribed to show changes at a sub-cellular level, making their importance in the assessment of physiological change even greater (Tom et al., 2004).

Besides the use of previously mentioned biomarker transcripts, endocrine linked genes may also be used to evaluate changes caused by toxicity in the environment (Devlin and Nagahama, 2002). Endocrine genes have the specific ability to show changes in toxicity caused by endocrine disrupting chemicals (EDCs) within the environment, and are particularly useful when studying EDCs like DDT (Santos et al., 2014). Cytochrome 1A (CYP1A) has previously been mapped in the Zebrafish Danio rerio (Trant et al., 2001), and these gene markers can be used as a guideline in this study as S. zambezensis CYP genes have not yet been fully mapped. Cytochrome genes play an important role in the detoxification process which takes place in the liver of organisms. Changes in the makeup of this enzyme can give an indication of toxicogenomic changes which have taken place in adult fish due to exposure to 4,4’-DDT.

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Chapter 3: Focus species

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3. Introduction More than 60% of the protein intake of adults in rural areas throughout Africa, is made up of fish and fish products (Adeleye, 1992). Fish are widely accepted as a delicacy in many countries regardless of economic, religious affiliation and education (Adeleye, 1992). Since it is a major source of protein, the handling, processing and harvesting of fish provides a means of financial support to millions of people. Providing 14% of all animal protein on a global scale, fish is the most important dietary animal protein in the world (Abolagba and Melle, 2008). Although fish is considered to be a healthier protein option because of the high content of long chain polyunsaturated fatty acids (LCPUFA’s) (Kabahenda et al., 2009), recent studies on their potential to bio-accumulate OCPs may mean that human populations are taking in much higher concentrations of xenobiotics than they realise (Coetzee et al., 2015).

Synodontis zambezensis (Peters, 1852) belongs to the family Mochokidae which is the largest family of Catfish in Africa, consisting of approximately 170 species (Skelton, 2001). The large number of species found within this genus means that they are the 3rd largest group of freshwater fish in Africa, surpassed only by Barbus and Haplochromis (Tweddle et al., 1979). Fish from this family can be difficult to identify because of their wide variety of characteristics. These fish are known as squeakers because of their ability to make stridulatory noises, caused by the rubbing of their pectoral fin spines against their bony head plates (Koblmüller et al., 2006). Their small to medium size combined with their attractive markings and colour morphs make them sought after freshwater aquarium species. Some species swim upside down and this has earned them their other common name, the upside-down catfish (Friel and Vigliotta, 2006). Another species belonging to the same family, Synodontis multipunctatus, is a brood parasitic catfish which relies on other fish species to look after and raise their young (Koblmüller et al., 2006), and yet another species displays Müllerian mimicry (Wright, 2011).

The reason this species was chosen for the study is because of its importance as a food source in many African countries, including Benin (Baras and Laèyè, 2003), Tanzania and South Africa (Coetzee et al., 2015; Daget et al., 1991). A survey done in the communities of the lower Phongolo River and floodplain, South Africa which included 521 people, showed that 64% of the interviewees actively caught and consumed this species (Coetzee et al., 2015). This species is eaten in this

30 | P a g e community because of their cartilaginous skeletal structure and forms an important protein source in the diets of infants. The fish are cooked by boiling or frying and then given to children (Smit et al., 2016). There is no large amount of bony material present within the skeleton of this fish, except within the head plate and the spines, making them easy to consume. Synodontis zambezensis are also very fatty fish and therefore supply essential fatty acids to young children (Coetzee et al., 2015). The skin of S. zambezensis is smooth and contains no scales, which also makes consumption much easier for smaller children (Merron et al., 1993).

3.1 Taxonomy Taxonomy can be defined as various things depending on the field of study it is being applied to, but the basic principles of recognising, classifying and placing the individual into a broader context are maintained throughout biological fields (Schuh, 2000). Taxonomy and systematics are terms, which are often used interchangeably (Schuh, 2000), but both systems have their shortcomings and often misrepresent species based on unreliable prominent features (Bruwer and Van der Bank, 2003). More recently, classification of fish species has been completed with a combination of morphological studies and biochemical genetics (Bruwer and van der Bank, 2002).

Kingdom: Animalia

Phylum: Chordata

Class: Actinopterygii (ray-finned fish)

Order: Siluriformes (catfish)

Suborder: Siluroidei

Family Mocholidae (squeakers, upside-down catfishes)

Genus and Species: Synodontis zambezensis (Peters, 1852; derived from the Zambezi River)

Common Names:

Brown squeaker (South Africa, Zimbabwe, USA, Zambia)

Bruin skreeubaber (South Africa)

Plain squeaker, Korokoro (USA)

Kanvaka, Kavuba, Kuya liwe, Kakolokotu, Mboma, Kahulu, Tshingondola (Angola)

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Nkholokolo (Malawi)

Ngonte (Tanzania)

3.2 External anatomy This species has a complex mouth with strong pectoral and dorsal spines as part of their fins (Skelton, 2001). Their pectoral spines are used predominantly for protection from possible predators (Willoughby, 1994) but also for communication through stridulatory noise production (Koblmüller et al., 2006). They are small to medium in size, have compressed bodies with a large adipose fin anterodorsally to their forked caudal fin (Figure 3.1) (Willoughby, 1994). They have well-developed cephalo-nuchal shields on the top and side of their heads, mainly used for protection (Willoughby, 1994).

Figure 3. 1: Image of Synodontis sp. adapted from Skelton (2001).

Morphological keys for Synodontis sp. identification have historically relied on four main characteristics; i) barbel morphology, ii) humeral process shape, iii) colour pattern and iv) moveable mandible tooth number (Friel and Vigliotta, 1994). Revisions of these identification keys have now included more descriptive characteristics that are able to distinguish between the subtle differences that exist among the different species. These additions to the identification key include; i) size, ii) shape and arrangement of premaxillary teeth, iii) size of caudal fin fork and iv) two measurements between the cranium and pectoral girdle (Bruwer and Van der Bank, 2003; Friel and Vigliotta, 2006).

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Synodontis sp. is a gonochoristic species which form distinct male and female individuals during development (Wijaya et al., 2017). The formation of males versus females is affected by environmental factors such as temperature (Devlin and Nagahama, 2002). The separate sexes are easily identifiable when they become sexually mature, and much like other catfish species, the males have a visible papilla with a pointed tip (Figure 3.2B). When the males are mature, the genital papilla often has a red tip (Friel and Vigliotta, 2006) and the female has a rounded papilla (Friel and Vigliotta, 2006) (Figure 3.2A).

Figure 3. 2: External sex differentiation of female (A) and male (B) Synodontis zambezensis indicating the anal orifice and the genital papillae.

3.3 Distribution The genus Synodontis has a wide distribution in Africa (Figure 3.3) and is found from the Lake Rukwa in Tanzania to the Phongolo River in South Africa (Daget et al., 1991). They are present within reaches of rivers where the water is slow flowing and are active during the late afternoon and evening hours making them almost completely nocturnal (Skelton, 2001).

Figure 3.3 shows the distribution of the species according to the International Union for Conservation of Nature (IUCN) (grey highlighting) as well as the countries in which literature has been published on this genus (green). The major rivers are indicated in blue and the assumption is that all of the countries through which these rivers flow will have a presence of Synodontis sp. but this is not indicated in any research conducted so far.

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Figure 3. 3: Distribution of Synodontis zambezensis within Africa. Grey highlighting indicates IUCN distribution data and green highlighted countries indicate published data sets on this species (Bills et al., 2010).

3.4 Feeding habits Synodontis sp. are considered to be scavengers, feeding on detritus, but have also been classed as omnivorous generalists (Friel and Vigliotta, 2006) actively hunting small insects like moths and crickets. Other sources of food include molluscs, annelids and crustaceans that are native to the system they are found in, and this evidence further supports that they may hunt more than they scavenge (Sanyanga, 1998). Some species have also adapted to filter feeding (Lalèyè et al., 2006).

3.5 Artificial spawning and its importance The production of protein in the form of fish has resulted in an increase in aquaculture throughout the world (Gjedrem et al., 2012). The amount of fish supplied as food through conventional fishing techniques is now far outweighed by the aquaculture industry (Gjedrem et al., 2012). Breeding fish in captivity is not only

34 | P a g e important as a source of protein for many countries (Abolagba and Melle, 2008; Adeleye, 1992; Gjedrem et al., 2012), but also as biological indicator species (OECD, 2014, 1992). Fish have been used for many years to indicate the toxicity of substances such as pharmaceuticals, pesticides and industrial waste products (OECD, 2014). Fish are exposed to chemicals of interest at various developmental stages to indicate the toxicity along the population gradient (Braunbeck et al., 2015). This makes the ability to breed fish artificially very important, since testing can be done at the embryonic stage (Braunbeck et al., 2015), juvenile stage (OECD, 1992) as well as adult stage (OECD, 2014) of development. Breeding fish in captivity results in juveniles that have not been exposed at all to the external environment, but instead are completely free of any possible contaminants found in many natural water systems (Kaiser and Rouhani, 1999; Naigaga et al., 2011). Toxin free fish are vital for the completion of successful and accurate laboratory based toxicity experiments (OECD, 2014). The artificial breeding of S. zambezensis was completed as part of this study for use in juvenile exposure experiments mentioned later on as well as publication of the techniques used in Chapter 4: Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852).

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Chapter 4 : Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852)

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Abstract Synodontis sp. are an important source of food in many African countries including, Nigeria, South Africa, Egypt and Ghana. This is likely due to the genus being an excellent source of fat in fish-based diets. However, there is also increased accumulation of lipophilic pollutants which can be transferred to the human population. In South Africa, Synodontis zambezensis is an important food source in rural communities, particularly along the Phongolo River system. The aim of this study was to breed Synodontis zambezensis in captivity, adapting a variety of protocols previously used in artificial breeding of Clarias gariepinus, Synodontis petricola and Synodontis nigromaculatus. A synthetic gonadotropin-releasing hormone compound (Aquaspawn®) was used to induce reproductive development followed by manual fertilisation of stripped eggs using sperm from dissected testis of sacrificed males. The embryonic development was consistent with other fish species and followed the same embryonic developmental stages. Initial hatching occurred within 26 hours after fertilisation and survival rates of fry and subsequent fingerlings was as high as 80%. S. zambezensis has previously been used as a bio-indicator species in field studies. Successful artificial breeding and rearing of this species improves the ability to use this indigenous species as a test organism in laboratory based assessments of chemicals and bio-accumulation of harmful chemicals such as 1,1'-(2,2,2-trichloro-1,1-ethanediyl)bis(4-chlorobenzene) (4,4’-DDT).

Keywords: artificial breeding; embryonic development; Aquaspawn®; Synodontis sp.

4. Introduction There are over 131 different Synodontis species (Friel and Vigliotta, 2011) of which many are highly sought after in the aquarium fish trade. They often have very attractive colour patterns and markings, are small, and relatively easy to keep (Kaiser and Rouhani, 1999). Captive breeding and rearing protocols need to be developed and improved to meet the demand for these fish (Kaiser and Rouhani, 1999). Previous breeding studies have focused on S. petricola (Kaiser and Rouhani, 1999), a desirable species native to Lake Tanganyika (Coulter, 1991) where the typical exports are wild-caught fish.

Synodontis sp. are also an important food source in many African countries, including; Nigeria (Akombo et al., 2014), South Africa (Coetzee et al., 2015), Egypt (Bishai and Abu Gideiri, 1965; Halim and Guma’a, 1989) and Ghana (Ofori-Danson,

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1992). The high fat content of Synodontis sp. (Gaspar et al., 2015) makes them an excellent source of fat in fish based diets. Unfortunately this high fat content may also increase the transfer of accumulated lipophilic pollutants to human populations (Kidd et al., 2001; Smit et al., 2016).

The species of focus for this study is S. zambezensis, the Brown Squeaker which is not considered an important species in the aquarium fish trade as it is particularly unremarkable (Skelton, 2001). However, it is an important food source for rural communities in the lower Phongolo River and floodplain (Coetzee et al., 2015). Questionnaires, which focused specifically on the use of fish in the community, were completed by community members (Coetzee et al., 2015; Smit et al., 2016) and indicated that a large majority of the S. zambezensis caught, is provided to infants as food. These fish are used as infant food because once the bony head plate and spines are removed, the remaining cartilage and tissue is easy to consume (Merron et al., 1993).

No published data exists on the successful artificial breeding of this species. Establishing and publishing a protocol outlining the successful breeding of this fish will allow for further scientific investigation of the species itself as well as its use as an indigenous bio-monitoring species.

The aim of this study was to successfully breed S. zambezensis in captivity through artificial fertilisation and to note the embryonic development in the fertilised eggs. Identification of any factors that may affect further development and survival were also noted during this study.

4.1 Materials and methods Permits for Gauteng (import and release of fish into the aquarium at the University of Johannesburg-CPE3 000492) as well as Limpopo (export of fish-CPE4 000111) have been issued and ethical clearance has already been given for the procedural processes and handling of the fish species used in this study (University of Johannesburg, Faculty Ethics committee, Faculty of Science). Fish were housed according to the guidelines provided in the SANS 10386 document for use of animals in research (SANS, 2008).

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4.1.1 Water quality preferences S. zambezensis is listed as a species of “Least Concern” by the International Union for Conservation of Nature (IUCN) and inhabits water sources with a pH range between 6.5 and 7.8 (Skelton, 2001). Their tolerance to temperature is quite broad with a range of 18.5˚C to 25˚C (Skelton, 2001). Sufficient places of refuge were provided through the placement of PVC piping and cloth covering the tanks. The preferred habitat of this species requires places to hide during the day as they are nocturnal animals (Skelton, 2001).

4.1.2 Artificial breeding setup Two mature males and two mature females, of three years in age, were selected for each session of artificial breeding and rearing. The two males and two females were placed into a 100 L holding tank (50 cm x 50 cm x 40.7 cm) with a false bottom, sponge filter and PVC piping for refuge (Figure 4.1). The tank was maintained at a constant temperature of 25˚C and a day/night 12/12 hrs photoperiod.

Figure 4. 1: Adult fish holding tank (100 L) setup with false bottom, sponge filter, 25 W heater and PVC piping.

4.1.3 Aquaspawn® Injection, egg stripping and fertilisation A volume of 0.1 mL Aquaspawn® (Aquaculture Innovations, Grahamstown, South Africa) (Kaiser and Rouhani, 1999) was injected into the dorsolateral muscle tissue of the mature fish 24 hrs prior to manual fertilisation. The holding tanks were fitted with false bottoms made with 5 mm oyster mesh to prevent the adults from eating any eggs released during the 24 hr holding period. After 24 hrs the eggs from each female were stripped into a dry glass petri dish (Figure 4.2A). Once all of the eggs

39 | P a g e from the two females were collected, one male was sacrificed and the testes were dissected out. Dissected testes were placed into a dry glass Petri dish and gently cut open to release the semen. If the first male’s testes were mature or ripe, white in colour, turgid (Owiti and Dadzie, 1989) and contained enough semen for adequate fertilisation of the collected eggs, then the second male was not dissected. Sperm was added to the eggs, gently mixed using a feather and tank water was added to activate the sperm (Figure 4.2B). Mixing was conducted for 2 mins or more to ensure even distribution and increased fertilisation rates (Figure 4.2C). Manual mixing also reduced clumping of fertilised eggs. The fertilised eggs were then transferred to an egg tumbler.

Figure 4. 2: Artificial fertilisation process; A) stripped eggs placed into dry glass Petri dish, B) semen from testes added to eggs and mixed with feather C) water from hatching tank added and mixed with sperm and eggs using a feather.

4.1.4 Egg tumbling and holding tank Egg tumblers are used for rearing experimental fish (Buss et al., 1970; Poston, 1983) and to separate organisms from detritus (Koosman and Newburg, 1977). In this study an egg tumbler placed in a 3 L beaker (Figure 4.3) was also used to prevent potential fungal growth on the fertilised eggs, a problem which plagued initial attempts.

Further reduction of potential fungal growth was limited by using reconstituted water (fish medium) (IWQS, 1998), consisting of 50 L deionized water to which: 23.52 g calcium chloride (CaCl2), 9.86 g magnesium sulphate (MgSO4), 5.18 g sodium bicarbonate (NaHCO3), 0.46 g potassium chloride (KCl) was added. Fungus clear (Pond Medic, active ingredients acriflavine, methyleneoxide and aminoacridine) was

40 | P a g e also added to the hatching tank to further reduce fungal growth at a concentration of 1 mL per 50 L and 0.1 mL per 3 L.

Figure 4. 3: Do it yourself (DIY) Egg Tumbler in a 3 L beaker with Fungus Clear treated fish medium.

4.1.5 Larvae, fry and fingerling maintenance After initial development and hatching, the larvae were transferred to 15 L tanks containing fish medium and a sponge filter pre-treated with a commercially available fungal preventing treatment (Pond Medic Fungus Clear, 0.1 mL). Larvae were sorted into 30 to 50 larvae per tank to reduce stocking density. The larvae were left until their yolk sacs had been absorbed; any larvae that died were removed twice daily.

The fry were fed semi-floating dry food, with a balanced nutrient profile and high content of amino acids (Hikari® Tropical First Bites®, Kyorin Co.,LTD. Kamihata Fish Industry Group). The tanks were maintained at a temperature of 25˚C and a day/night 12/12 hrs photoperiod. The growth tanks (23.2 cm x 30.5 cm x 22.5 cm) were filled to a volume of 15 L, a stocking density of between two and five fry per litre (30 to 50 fry per growth tank) and covered in shade cloth since juveniles are photosensitive (Sautter et al., 2007).

The fingerlings were fed juvenile Daphnia pulex and Hikari® Tropical First Bites® ad lib, but not too much, to prevent fouling the water in the tanks, and the above mentioned tank conditions were maintained.

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4.1.6 Microscopic observation of egg development After successful fertilisation a few eggs were placed into a glass petri dish for observation under a dissecting microscope fitted with a 2.0 x 45 50 28 lens (Zeiss Stemi DV4 Stereo Microscope, ZEISS, South Africa). Photographs were taken at frequent time intervals to record the development of the fertilised eggs until hatching.

4.2 Results and discussion 4.2.1 Breeding The artificial breeding of S. zambezensis was successful and the survival rate of the juveniles (± 90%) is comparable to those in a study using S. petricola (Kaiser and Rouhani, 1999). Kaiser and Rouhani (1999) were more successful in breeding S. petricola “naturally” than in this study. The methods described by Kaiser and Rouhani (1999) were initially attempted for a 30 day period to allow for natural spawning behaviour in the holding tanks once the males and females had been injected with Aquaspawn®. These attempts were unsuccessful because the eggs that were released into the tanks were either unfertilised, or they were infected with fungus shortly after fertilisation. In order to increase the fertilisation, hatching and subsequent survival rate, a manual method of fertilisation was opted for. The fertilisation method in this study was adapted from the artificial breeding of C. garipienus (Abdulraheem et al., 2012; De Graaf and Janssen, 1996) and S. nigromaculatus (Van der Waal, 1986).

4.2.2 Embryonic development Fish embryos are meroblastic and develop through discoidal cleavage (Kimmel et al., 1995). S. zambezensis embryos show development comparable with that of S. nigromaculatus (Van der Waal, 1986) but the development seen in this study is much quicker. The embryos hatched within 26 hours after fertilisation (Table 4.1). The reason for the increased speed of development is not known.

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Table 4 1: Time after fertilisation, description of developmental stages and photographs of Synodontis zambezensis bred in captivity.

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4.3 Conclusions and recommendations S. zambezensis can be successfully bred in captivity for laboratory investigation. This species may be an appropriate alternative to use instead of larger catfish like C. gariepinus and their smaller size allows for higher numbers of fish to be placed into exposure tanks. Once they reach one month in age they are easy to care for and maintain within an aquarium environment. It is recommended that further studies into breeding preferences be conducted so that the sacrifice of a male may be avoided when attempting to breed this species in captivity. It would be desirable that the breeding follows a more natural route. Further investigation into the slow growth rates of this genus should also be considered.

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Chapter 5 : Baseline bio- accumulation concentrations and resulting oxidative stress in Synodontis zambezensis after an acute laboratory exposure to 4,4’-DDT

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Abstract The use of 1,1'-(2,2,2-trichloro-1,1-ethanediyl)bis(4-chlorobenzene) (DDT) as a pesticide for the control of insect vectors responsible for the spread of many life threatening diseases was officially banned in 1972 by the United States Environmental Protection Agency (USEPA). It was banned throughout the world, in most developed countries, because of the toxic effects it causes in wildlife, including birds and fish. However, DDT is still used in approximately 43 African countries, including South Africa, to control the spread of malaria. The lipophilic nature of DDT and therefore its persistence in the environment makes it extremely important for laboratory based studies to be conducted in an effort to evaluate the accumulation potential and possible physiological effects of DDT in aquatic organisms under controlled conditions. The aim of this study was to establish baseline bio- accumulation concentrations within Synodontis zambezensis following an acute exposure to 4,4’-DDT. The three metabolites analysed were 4,4’-DDE, 4,4’-DDD and 4,4’-DDT. None of the 2,4’-isomers were analysed in this study since the acute exposure used a solution of 98.7% pure 4,4’-DDT (Sigma-Aldrich PESTANAL®, Analytical Standard, CAS-No 50-29-3, Batch number SZBE057XV) and not a mixture of 4,4’-DDT and 2,4’-DDT as found in technical grade DDT. Soxhlet extraction of tissue samples and liquid/liquid extraction of water samples followed by analysis through Gas-chromatography mass-spectrophotometry was completed. Mean 4,4’- DDE, 4,4’-DDD and 4,4’-DDT concentrations ranged from 15.34 ng/g to 45.34 ng/g, 28.16 ng/g to 63.25 ng/g and 28.64 ng/g to 96.21 ng/g respectively. All of the accumulated concentrations fell within environmentally relevant concentrations with no input through the food web. The accumulated concentrations of 4,4’-DDT and its three metabolites resulted in oxidative stress responses within the gills and the liver tissue of S. zambezensis. Significant differences (p≤0.05) were observed between malondialdehyde (MDA) and reduced glutathione (GSH) within the liver and in superoxide dismutase (SOD), catalase (CAT) and reduced glutathione (GSH) in the gills.

Key words: 4,4’-DDT; Synodontis zambezensis; bio-accumulation; acute exposure; biomarkers; oxidative stress.

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5. Introduction The control of pests through the use of pesticides is a requirement in the agricultural sector which relies on the use of herbicides, fungicides and insecticides to increase high-quality production of desired crops (Bennett et al., 2003; Oerke, 2006). Increasing populations leads to an increasing demand for products including food and clothing, inevitably leading to an increase in the production of pesticides throughout the world (Van Dyk et al., 2007). Although many pesticides are designed to affect and control a specific group of organisms, many have the potential to cause harm to non-target organisms resulting in detrimental ecosystem changes (Davies et al., 2007).

1,1’-(2,2,2-trichloro-1,1-ethanediyl)bis(4-chlorobenzene) also known more commonly as 4,4’-DDT was originally synthesised in 1874 by the Austrian chemist, Othmar Zeidler for the use of pest control on crops of corn (Zeidler, 1874). Once its insecticidal properties were discovered in 1939 it was put into large-scale synthesis and application to eradicate the spread of diseases such as malaria, typhus and the bubonic plague (WHO, 1979). Although DDT was banned in 1972 by the United States Environmental Protection Agency (USEPA) it is still used in the north eastern parts of South Africa for the control of malaria transmission (Barnhoorn et al., 2009). DDT is applied through Indoor Residual Spraying (IRS) annually in the Limpopo Province, in particular, the Vhemba District Municipality (Urbach, 2007; WHO, 2002). Through the movement of spray drift from sites of application, accumulation takes place in aquatic systems, as well as human and animal tissues, in areas that may not have been sprayed with 4,4’-DDT (Aneck-Hahn et al., 2006; Gaspar et al., 2015; Whitworth et al., 2014). The accumulation of organochlorine pesticides (OCPs), a group which includes 4,4’-DDT, can cause toxic effects in fish, and eventually lead to disruption on a physiological and biological level (Da Cuña et al., 2013).

Monitoring potential changes within specific organisms has become increasingly important and this can be done through various means (Dix et al., 2007). The organisms used to study the effects of chemicals will differ depending on the aquatic system that is being monitored (Smolders et al., 2003a). Test organisms used in biomonitoring studies are known as biological indicators or bio-indicators (Naigaga et al., 2011) and have been defined as meeting the following criteria (Zhou et al., 2008): 1) organisms should accumulate high levels of pollutants without death, 2)

47 | P a g e organisms should represent the local population, 3) organisms should be widely distributed and have a high abundance for repetitive sampling, 4) organisms should have a long lifespan, 5) organisms should be easy to sample, 6) organisms should live in water (aquatic) and 7) organisms should occupy an important position in the food chain.

Bio-monitoring with the use of fish as a bio-indicator is considered a successful means of monitoring the current state of a system as native fish species are constantly exposed to changes which occur within the system (Van der Oost et al., 2003). The species used for this study is Synodontis zambezensis, commonly known as the brown squeaker due to the stridulatory “squeaking” noise it makes when it is stressed (Koblmüller et al., 2006). This species is an important food source for young children in the local communities surrounding the lower Phongolo River and floodplain (Coetzee et al., 2015). This species is known for its high-fat content and may, therefore, bio-accumulate pesticides to much higher concentrations than those seen in other fish species (Coetzee et al., 2015; Smit et al., 2016). This poses an increased threat to human populations relying on this species as a source of nourishment (Coetzee et al., 2015).

Biomarkers of effect, which reflect oxidative stress in the cells and tissue of organisms, were chosen for analysis in this study (Almroth et al., 2008). Oxidative stress as defined by Almroth et al. (2008) refers to an imbalance in pro-oxidants and antioxidants within an organism which may lead to damage. Reactive oxygen species (ROSs) form within the tissues of aquatic organisms in response to exposure to xenobiotics (Ferreira et al., 2005). The three main ROSs that form within fish species and result in oxidative stress are hydrogen peroxide, superoxide radicals and hydroxyl radicals (Atli et al., 2006). Superoxide dismutase (SOD) is an antioxidant which forms in response to the breakdown of fatty components of cell membranes by the action of ROSs (Rola et al., 2012). Catalase (CAT) works together with SOD to neutralize the resulting ROSs caused by aerobic metabolism or exposure to pollutants (Khessiba et al., 2005). The formation of the ROS malondialdehyde (MDA) following the oxidation of polyunsaturated fatty acids is an indication of oxidative stress and is measured through the biomarker MDA (Flohr et al., 2012). Structural changes, which occur in amino acid groups, result in damaged proteins known as protein carbonyls (PC) (Zusterzeel et al., 2001). The formation of

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PC is irreversible and is used as a biomarker of oxidative stress (Almroth et al., 2008, 2005). Glutathione (GSH) acts on various ROSs to prevent damage to components present within the cell of affected organisms (Filipak Neto et al., 2008). This antioxidant acts directly on ROSs formed through lipid and protein peroxidation, forming oxidized glutathione disulphide (GSSG) (Slaninova et al., 2009).

The aim of this study was to establish baseline accumulation data within S. zambezensis, which may reflect the accumulation that would take place after application of 4,4’-DDT through IRS in nearby areas or after a spill event. The possible effects of such an exposure will be reflected by oxidative stress biomarkers, resulting in an assessment of the stress experienced by these individuals due to an acute exposure to 4,4’-DDT.

5.1 Materials and methods 5.1.1 Fish collection Fish were collected from an aquaculture facility in Limpopo Province and transported back to the University of Johannesburg in a 1000 L plastic transport tank. The addition of oxygen to the tank through compressed air pumps, aided in the reduction of mortality during transport (Mlambo et al., 2009).

5.1.2 Acclimation At the University of Johannesburg, fish were acclimatised for two months before any exposure experiments were performed. The environmental room in which the fish were housed was kept at a constant temperature of 26˚C and a day/night cycle of 12/12 hours. Borehole water was circulated from the reservoir through the biological filter and pumped back into the reservoir (Mlambo et al., 2009). Feeding of the fish did not take place during the first 72 hours after arrival as this can prove lethal when fish are experiencing stress (Barton, 2002).

Sufficient places of refuge were provided and shade cloth was placed over the tanks since this species is light sensitive (Sautter et al., 2007). The preferred habitat of this species requires places to hide during the day as they are nocturnal animals (Skelton, 2001). Fish were housed according to the guidelines provided in the SANS 10386 document for use of animals in research (SANS, 2008).

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5.1.3 Exposure concentrations The stock solution used for the exposure experiments was prepared by dissolving 4,4’-DDT (98.7%, Sigma-Aldrich PESTANAL®, Analytical Standard, CAS-No 50-29- 3, Batch number SZBE057XV) in ethanol (95%, Merck Millipore). Ethanol was used as the organic solvent for the 4,4’-DDT because it is insoluble in water and is therefore required to be dissolved in an organic solvent before addition to any water source (Pontolillo and Eganhouse, 2001).

One hundred milligrams 4,4’-DDT was dissolved in 1 L ethanol to make up a stock solution with a concentration of 98 700 ng/mL. From this stock solution, concentrations to be used for the 96-hour exposure period were calculated based on toxicological information gathered from the Material Safety Data Sheet (according to regulations (EC) No. 1907/2006, Version 5.2 Revision Date 26.11.2014) and environmentally relevant concentrations (Smit et al., 2016).

The concentrations chosen were intended to correspond with environmentally relevant concentrations, but they were not intended to cause mortality in the exposed fish. The highest exposure concentration was six times higher than the LC50 values found in Pimephales promelas and Lepomis macrochirus juveniles (Sigma-Aldrich, 2014). This exposure was a sub lethal exposure to 4,4’-DDT and was intended to allow for the determination of baseline bio-accumulation and resulting oxidative stress biomarker concentrations. 2,4’-DDT and its degradation products were not considered in this study.

The exposure was run in duplicate rather than in triplicate because of the amount of space required for the exposure as well as the limited number of fish available. The rest of the adult fish collected from the aquaculture unit in Limpopo were intended for experimentation in different sections of the study.

5.1.4 Exposure system and transfer of fish Each of the 16 exposure tanks was 100 L in volume. Exactly 50 L of borehole water was added to each tank and allowed to stand with corner filters for two weeks. After the two weeks establishing period, five fish from the main holding tank were added to each of the 16 exposure tanks. The fish were left in the exposure tanks for a further two week acclimation period during which they were fed once every three days. Feeding stopped 24 hours before the commencement of the exposure.

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Day 1: 32 mL ethanol was added to the solvent control (tank 2:1 and 2:2) and the first two 4,4’-DDT volumes of stock solution were added to the respective tanks 3:1 and 3:2 = 1 mL stock solution resulting in 1.974 ng/mL concentration in exposure tanks; 4:1 and 4:2 = 2 mL stock solution resulting in 3.948 ng/mL concentration in exposure tanks.

Day 2: Volumes of the stock solution were added to the remaining eight tanks 5:1 and 5:2 = 4 mL stock solution resulting in 7.896 ng/mL concentration in the exposure tanks; 6:1 and 6:2 = 8 mL stock solution resulting in 15.792 ng/mL concentration in the exposure tanks; 7:1 and 7:2 = 16 mL stock solution resulting in 31.584ng/mL concentration in the exposure tanks; 8:1 and 8:2 = 32 mL stock solution resulting in 63.168 ng/L concentration in the exposure tanks.

The concentrations were added one hour apart on both days to allow for adequate dissection times of five fish per hour.

5.1.5 Water quality Standard in situ water quality parameters, dissolved oxygen (DO), oxygen percentage (O2%), total dissolved solids (TDS), electrical conductivity (EC), temperature (˚C) and pH, were measured in each tank at 24-hour intervals. The first readings were taken 20 mins after the addition of the relevant concentrations to allow for a homogenous mixture of water and 4,4’-DDT stock solution or 95% ethanol in the case of the solvent control. The water parameters were measured using a Eutech Multi-Parameter (PCTestr™ 35) instrument.

5.1.6 Tissue sample collection Biometric data of each fish was taken before and during dissection. The standard and total lengths were measured. Total mass, gutted mass, liver mass and gonadal mass were all recorded using a Sartorius Basic BA210S scale.

Muscle tissue was excised from the lateral side of each fish for bio-accumulation analysis. This tissue was wrapped separately in tinfoil and placed into individual zip- lock bags. Liver and gill tissue was removed from each fish for biomarker analysis and placed into separate falcon tubes containing Henriksson’s stabilising buffer (Henriksson et al., 1986). All samples were frozen at -80˚C until further analysis.

5.1.7 Water sample collection Water samples were collected in brown glass bottles from each tank at 24-hour intervals. The first water sample was collected half an hour after the addition of the

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4,4’-DDT stock solution to allow for homogeneity of the water and 4,4’-DDT stock solution. 5.1.8 DDT analysis of tissue and water Tissue extraction The procedure of Yohannes et al.,(2013) was followed for the extraction and analysis of 4,4’-DDT and its degradation products from muscle tissue samples. One modification was made to this protocol, but the standards used remained the same as in the original extraction technique. The modification was a shorter extraction time, from six hours to four hours and thirty mins. This adjustment was made with the introduction of updated system software. The organic solvents used during extraction were acetone (≥99.8%, Sigma-Aldrich, 34850) and hexane (≥97.0%, Sigma-Aldrich, 34859) at a ratio of 3:1 and the organic solvents used for extract clean-up were dichloromethane (DCM) (≥99.8%, Sigma-Aldrich, 34856) and hexane at a ratio of 2:7. A wet mass of 10 g of muscle tissue was homogenized in anhydrous sodium sulphate (≥99.0%, Sigma-Aldrich, 239313) and was placed into a pre-washed (acetone and hexane) extraction thimble. The surrogate standard 2,4,5,6-tetrachloro- m-xylene (TCmX) (200 µg/L in methanol, Sigma-Aldrich, 48317) was added to the homogenate and the extraction thimble was placed into the extraction chamber with 150 mL of the extraction solvent. The extraction was completed with Soxtherm Apparatus (S306AK Automatic Extractor, Gerhardt Germany). The extract was concentrated to ±2 mL using a rotary evaporator and diluted to 10 mL with hexane. Twenty percent of this mixture was used to determine the fat content of the fish through gravimetric lipid determination and the remaining 80% was dried to half of its volume with nitrogen gas and subjected to a clean-up procedure to remove excess fat. Due to the high-fat content, a double clean-up procedure was required. The first clean-up was done using gel permeation chromatography (G-Prep GPC8100, GL Sciences, Japan) with a 1:1 mixture of DCM and hexane. The resulting solution was concentrated again to ±2 mL and subjected to a second clean-up procedure. The second clean-up was done using 4 g activated Florisil® (≥80%, Sigma-Aldrich, 03286), a small amount of anhydrous sodium sulphate and glass wool (Sigma- Aldrich, 18421). One hundred and twenty mL of a 1:1 mixture of DCM:hexane was allowed to flow through the column in a drop wise manner. The resulting solution was again concentrated with a rotary evaporator, and further dried with nitrogen gas.

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The final liquid was reconstituted in 100 μL of n-decane (≥95%, Sigma-Aldrich, 30570) and analysed on a Shimadzu QP2010 GC-MS.

Water extraction Extraction of 4,4’-DDT and its degradation products from the water samples taken every 24 hrs was done according to the extraction methods outlined in Mahugija et al. (2015) with some modifications. The modifications included: smaller volumes of water collected for extraction and the use of clean-up with activated Florisil® instead of silica gel or alumina. Each 250 mL sample was collected in a brown glass bottle. These samples were frozen at -20ºC until extraction and analysis took place.

Each sample was extracted by first placing the sample into a Squibb style separating funnel and adding 60 mL DCM, the mixture was vigorously shaken for 30 secs and the organic layer (bottom layer) was drained into an Erlenmeyer flask. This was done three times, resulting in 180 mL DCM extract. Anhydrous sodium sulphate was added to this extract and mixed together in order to remove any excess water from the mixture. This “dry” extract was filtered through filter paper, glass wool and more anhydrous sodium sulphate into a flat-bottomed rotary flask and concentrated to ±0.5 mL using a rotary evaporator. The resulting evaporated sample was spiked with 100 µL of 100 ppb internal standard 3,3’,4,4’-tetrachlorobiphenyl (PCB #77, Sigma- Aldrich, 35496) before the clean-up of the sample. The spiked samples were cleaned up through glass wool, 3 g of activated Florisil® and 1 g of anhydrous sodium sulphate after the addition of 100 mL of hexane: DCM (1:1) clean-up solvent. The solution was collected in a second flat bottomed evaporation flask and evaporated to ±1 mL. The resulting residue was concentrated under a light stream of nitrogen gas to near dryness and finally reconstituted in 100 μL of n-decane for later analysis using a Shimadzu QP2010 GC-MS add (.

Gas chromatography-mass spectrometry Gas chromatography-mass spectrophotometry (GC-MS) (Shimadzu QP2010 GC- MS) was used to quantify the amount of 4,4’-DDT and its degradation products in the prepared tissue and water samples. Calibration curves were used to quantify the amount of 4,4’-DDT and its potential degradation products, 4,4’-DDE and 4,4’-DDD.

A Shimadzu 2010 plus GC-MS with AOC 20i autosampler was used. Instrument parameters were based on EPA Method 8081b (USEPA, 2007). A Zebron ZB- 5Msplus capillary column (30 m x 0.25 mm i.d., 0.25 μm film thickness) was used.

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Splitless injections of 1 μL were performed. The temperature of the GC oven was kept constant at 60˚C for 2 mins, after which it was ramped up by 10˚C/min where it was kept constant for 2 mins. The signal was detected using electron ionization mass spectrometry (EI-MS) with and interface temperature of 280˚C and ion source at 230˚C. Selected Ion Monitoring (SIM) was used for the analysis since the concentrations of 4,4’-DDT and its degradation products were very low. The following m/z values were selected for SIM: 246, 248, 176, 318, 235, 167. Each sample was analysed in triplicate and the mean of each sample was taken as the final concentration.

In order to determine the percentage of the initial dose of 4,4’-DDT accumulated by fish and extracted from the water, the presence of degradation products had to be taken into account. During degradation/metabolisim a stoichiometric ratio of 1:1 for 4,4’-DDT and its respective degradation products exists. Therefore the masses obtained from the calibration experiments (GC-MS) were converted to moles. The moles of 4,4’-DDT, 4,4’-DDE and 4,4’-DDD were then added to get the total number of moles of 4,4’-DDT from which it originated. The total number of moles were then converted to mass to give 4,4’-DDT equivalents.

Therefore, DDT equivalents were determined using the following equation:

퐷퐷푇 푒푞푢푖푣푎푙푒푛푡푠 푖푛 푛푔/푔 푥 푛푔 퐷퐷푇 1 푚표푙푒 퐷퐷푇 푦 푛푔 퐷퐷퐷 1 푚표푙푒 퐷퐷퐷 = [( × ) + ( × ) 푔 푠푎푚푝푙푒 354.48 푔 퐷퐷푇 푔 푠푎푚푝푙푒 320.04 푔 퐷퐷퐷 푧 푛푔 퐷퐷퐸 1 푚표푙푒 퐷퐷퐸 354.48 푔 퐷퐷푇 + ( × )] × 푔 푠푎푚푝푙푒 318.03 푔 퐷퐷퐸 1 푚표푙푒 퐷퐷푇

where x, y and z represent the values obtained from calibration experiments (i.e. GC-MS) for 4,4’-DDT, 4,4’-DDD and 4,4’-DDE respectively (Table 4).

Quality control and quality assurance The internal standard used for the water samples (PCB#77) was added at a volume of 100 μL and a concentration of 100 ppb before the clean-up procedure. The internal standard used for the tissue samples was TCmX. All recoveries of the internal standard were acceptable based on previous literature (Yohannes et al., 2013) and were above 85%. 4,4’-DDT and all possible degradation products were identified using Selected Ion Monitoring (SIM) and retention times were checked and

54 | P a g e assigned to each peak. Multi-level calibration curves were created for each metabolite and linearity was successfully achieved (R2≥0.995). Procedural and spiked blanks were analysed along with samples to assure quality control according to literature and reported results showed acceptable recoveries between 90% and 105% (Yohannes et al., 2014). The concentrations of tissue samples are expressed in lipid weight. The concentrations for water samples are however expressed in wet weight.

5.1.9 Biomarker analysis of muscle and gill tissue The biomarker protocols outlined in the section to follow were all completed within standard operating guidelines outlined in the relevant literature. All protocols used a 96 well microtiter plate and protein content was determined in the tissues for each of the biomarkers according to the method of Bradford (1976). Absorbance was measured at 630 nm with bovine serum albumin (BSA) used as a standard. All biomarker activities are expressed per milligram protein to allow for standardisation.

Superoxide dismutase The protocol applied for the analysis of SOD was adapted from Greenwald (1985). Zero point two grams of liver tissue was homogenised in 200 µL General Homogenising Buffer (GHB) and the homogenate was centrifuged (High Speed Refrigerated Centrifuge: NovaFuge B113-21R, Senova Biotech Co.,Ltd) at 7393 g. The reaction was initiated by adding 25 µL 1,2,3-benzenetriol (Pyrogallol ≥98%, Sigma-Aldrich) solution just before analysis. The measurements were recorded for 10 mins total length with excitation readings being taken at 60 sec intervals. The samples were analysed on a Multi-Detection microplate reader (FLx800 Fluorescence Microplate Reader, BioTek® Instruments, Inc.). Analysis was carried out in a dark room as pyrogallol is extremely light sensitive.

Catalase The procedural methodology for CAT activity was adapted from Cohen et al. (1970). Phosphate buffer (1 mL, 0.01 M, pH 7) and 0.1 g of liver tissue were homogenised and centrifuged at 5590 g. for 10 mins at 4°C. The amount of unreacted KMnO4 was then measured spectrophotometrically at 409 nm using an automated microplate reader (Elx800-Universal Microplate Reader, BioTek® Instruments, Inc.).

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Malondialdehyde Malondialdehyde (MDA) was analysed through procedures and methodologies adapted from Ohkawa et al. (1979) including modifications by Üner et al. (2006). Homogenate was prepared by homogenising 0.05 g of liver tissue in 250 µL of Tris/sucrose buffer and then centrifuging at 5045 g. for 10 mins at 4°C. The samples and blank were added in triplicate to a microtiter plate at a volume of 245 µL and analysed at 540 nm using an automated microplate reader (Elx800-Universal microplate reader, BioTek® Instruments, Inc.).

Protein carbonyl Methodologies were adapted from Parvez and Raisuddin (2005) as assayed by Levine et al. (1990) and also contains modifications applied by Floor and Wetzel (1998). Zero point one grams of liver tissue was homogenised in GHB and centrifuged at 10 188 g. for 30 mins at 4°C. After the above mentioned protocols were followed, 100 µL of the resulting liquid was analysed at 366 nm using an automated microplate reader (Elx800-Universal microplate reader, BioTek® Instruments, Inc.).

Reduced glutathione Reduced glutathione (GSH) content was determined following the protocols outlined by Cohn and Lyle (1966). A homogenate of 100 mg of tissue and 250 µL Tris/sucrose buffer was processed following the above mentioned protocol and 100 µL of the sample was analysed at 420 nm emission with excitation of 350 nm (Elx800-Universal microplate reader, BioTek® Instruments, Inc.).

5.1.10 Statistics SPSS version 25 (IBM Software Group) was used for one-way analysis of variance (ANOVA) following log transformation of all accumulation and biomarker data to ensure homogeneity. Significant differences (p≤0.05) between concentrations and skewness of data were determined using the Tukey Post Hoc test (Yohannes et al., 2013). Multivariate statistics were applied to the bio-accumulation and biomarker data. Constrained redundancy analysis (RDA) was applied to determine the influence of 4,4’-DDT and its degradation products on the spread of the oxidative stress biomarkers analysed (Quinn and Keough, 2002). The constrained RDA was completed with the use of Canoco 5 software.

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5.2. Results 5.2.1 Biometric data The standard length of the exposed fish ranged from 69 mm to 106 mm with the total lengths ranging from 89 mm to 132 mm (Table 5.1). The total mass ranged from 7.22 g to 23.82 g with no significant differences (p≤0.05) seen in the lipid contents of the exposed fish. The gonadal and liver mass ranged from 0.02 g to 1.40 g and 0.06 g to 0.26 g respectively. These organ parameters where measured to calculate the Gonadosomatic and Hepatosomatic Indices, but these indices yielded no significant differences (p≤0.05) between fish in the varying exposure concentrations.

Table 5. 1: Biometric data (Mean ± SD) and range of Synodontis zambezensis exposed to 4,4’-DDT during an acute exposure. Exposure Number of Standard Total Length Gonadal Mass Total Mass (g) Liver Mass (g) Lipid % Conc. (ng/mL) Fish (n) Length (mm) (mm) (g) 91.50 ± 3.07 112.8 ± 10.26 14.83 ± 3.66 0.139 ± 0.04 0.20 ± 0.25 7.07 ± 0.88 Control 10 (79 – 106) (96 – 131) (9.23 – 22.35) (0.07 – 0.19) (0.05 – 0.86) (4.58 – 10.19) Solvent 82.10 ± 7.26 104.90 ± 8.80 12.41 ± 3.27 0.16 ± 0.06 0.14 ± 0.14 10.25 ± 0.35 10 Control (71 – 97) (90 – 121) (7.36 – 17.70) (0.08 – 0.26) (0.02 – 0.52) (9.22 – 11.51) 85.70 ± 6.11 108.10 ± 7.65 12.84 ± 2.97 0.11 ± 0.03 0.09 ± 0.06 9.03 ± 1.10 1.97 10 (79 – 96) (102 – 124) (9.68 – 18.19) (0.09 – 0.19) (0.04 – 24) (6.94 – 11.56) 82.30 ± 7.62 104.00 ± 7.29 11.54 ± 3.18 0.10 ± 0.03 0.09 ± 0.11 6.87 ± 1.42 3.95 10 (71 – 99) (92 – 119) (8.03 – 18.72) (0.06 – 0.16) (0.02 – 0.40) (2.73 – 10.12) 88.70 ± 10.12 113.20 ± 12.30 14.54 ± 4.75 0.14 ± 0.06 0.25 ± 0.28 4.62 ± 0.41 7.90 10 (69 – 105) (89 – 132) (7.22 – 23.82) (0.06 – 0.24) (0.03 – 1.00) (3.60 – 6.21) 88.10 ± 8.12 110.80 ± 9.86 14.05 ± 3.60 0.15 ± 0.04 0.23 ± 0.18 5.55 ± 0.30 15.79 10 (78 – 102) (98 – 131) (9.55 – 21.47) (0.10 – 0.24) (0.07 – 0.56) (5.32 – 6.56) 87.50 ± 6.10 113.70 ± 8.59 14.62 ± 3.99 0.17 ± 0.06 0.31 ± 0.42 4.30 ± 0.29 31.58 10 (78 – 95) (101 – 124) (9.43 – 22.08) (0.07 – 0.27) (0.02 – 1.40) (3.34 – 5.19) 87.60 ± 7.77 107.60 ± 11.10 12.83 ± 3.96 0.16 ± 0.05 0.18 ± 0.24 5.13 ± 1.03 63.17 10 (81 – 100) (99 – 126) (8.90 – 19.16) (0.12 – 0.24) (0.03 – 0.60) (1.97 – 8.61)

5.2.2 Water quality The water quality parameters changed slightly during the acute exposure period but were fairly constant (Table 5.2). The mean pH readings across all 16 exposure tanks ranged from 7.45 to 7.79 which falls within the pH tolerance of Synodontis zambezensis (Skelton, 2001). The dissolved oxygen (DO mg/L) and oxygen

percentage (O2%) was very low during the initial measurement taken 30 mins after the addition of the concentrations and the ethanol in the case of the solvent control. The oxygen levels stabilized hereafter and resulted in a mean across all tanks

ranging from 5.59 mg/L to 6.70 mg/L (DO) and 74.50% to 84.49% (O2%). Since the exposure tanks were kept in an environmental room maintained at a constant

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temperature of 26˚C, the mean temperature of the water did not fluctuate much throughout the exposure period and ranged from 25.41˚C to 26.17˚C, again falling within the temperature tolerance range of Synodontis zambezensis (Skelton, 2001). Increases in conductivity (EC) and total dissolved solids (TDS) are directly proportional to increases in the concentration of 4,4’-DDT added to the exposure tanks. The mean EC and TDS ranged from 263.10 µS/cm to 1269.40 µS/cm and 133.20 mg/L to 649.50 mg/L respectively.

Table 5. 2: In situ water quality parameters (mean ± SD) and range readings taken of the exposure tanks during the course of the 96-hour 4,4’-DDT exposure experiment. pH, dissolved oxygen (DO), oxygen percentage (O2%), electrical conductivity (EC), temperature (˚C) and total dissolved solids (TDS). Exposure Conc. pH D.O (mg/L) O2 (%) EC (µS/cm) Temp. (°C) TDS (mg/L) (ng/mL) 7.51 ± 0.17 6.13 ± 0.76 75.51 ± 9.03 263.10 ± 1.63 26.04 ± 0.33 133.20 ± 0.69 Control (6.92 – 7.93) (3.81 – 7.69) (48.20 – 94.60) (259.00 – 268) (25.40 – 28.50) (131.00 – 135.00) 7.59 ± 0.22 6.05 ± 0.80 74.50 ± 9.70 785.20 ± 17.00 25.78 ± 0.10 394.00 ± 8.14 Solvent Control (6.61 – 7.98) (3.76 – 8.86) (46.60 – 108.20) (660.00 – 757.00) (25.3 – 26.1) (370.00 – 436.00)

1.97 7.59 ± 0.21 5.69 ± 0.48 70.31 ± 5.85 692.00 ± 12.19 25.93 ± 0.12 344.40 ± 6.57 (6.63 – 7.91) (4.27 – 6.64) (53.10 – 82.30) (659.00 – 757.00) (25.40 – 26.30) (328.00 – 379.00)

3.95 7.45 ± 0.21 5.60 ± 0.57 69.33 ± 7.03 962.70 ± 14.77 26.17 ± 0.11 482.20 ± 7.40 (7.12 – 8.03) (5.06 – 8.63) (61.80 – 105.20) (1125.00 – 1139.00) (25.30 – 25.80) (560.00 – 618.00)

7.90 7.64 ± 0.17 6.91 ± 0.55 84.49 ± 6.78 1160.20 ± 14.43 25.41 ± 0.08 588.80 ± 10.92 (7.12 – 8.03) (5.06 – 8.63) (61.80 – 105.20) (1125.00 – 1235.00) (25.30 – 25.80) (560.00 – 618.00)

15.79 7.79 ± 0.12 6.70 ± 0.55 82.21 ± 6.67 1269.40 ± 38.97 25.67 ± 0.09 633.70 ± 19.66 (7.36 – 8.00) (4.84 – 8.56) (59.50 – 104.00) (1113.00 – 1499.00) (25.20 – 26.10) (557.00 – 748.00)

31.58 7.61 ± 0.15 6.12 ± 0.36 75.39 ± 4.48 1300.80 ± 9.51 25.71 ± 0.18 649.50 ± 4.50 (7.03 – 7.95) (4.80 – 6.88) (59.20 – 83.70) (1254.00 – 1359.00) (25.30 – 26.30) (629.00 – 678.00)

63.17 7.60 ± 0.21 6.19 ± 0.43 76.36 ± 5.19 1139.80 ± 10.11 26.01 ± 0.09 568.70 ± 5.06 (6.31 – 8.02) (4.72 – 7.39) (58.40 – 90.60) (1101.00 – 1186.00) (25.60 – 26.30) (536.00 – 591.00)

5.2.3 Levels of 4,4’-DDT in Synodontis zambezensis All three of the degradation products were found in the fish exposed to the 4,4’-DDT in all six exposure concentrations. The 4,4’-DDT used was of 98.7% purity purchased from Sigma-Aldrich PESTANAL® as an Analytical Standard (CAS-No 50- 29-3, Batch number SZBE057XV). The readings are reported in lipid mass because there were no significant differences (p≤0.05) found in the fat content of the fish in each of the exposure tanks (Table 5.1). Accumulated concentrations for the control tanks are not included in the results section because, as expected, there was no 4,4’-DDT or its degradation products found in the tissue of the fish in these tanks. The S. zambezensis used in the exposure were bred and grown completely in captivity before the exposure took place. Accumulated concentrations varied

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throughout the exposure tanks and did not always reflect the nominal concentrations added at the beginning of the 96-hr exposure period (Table 5.3). The mean 4,4’-DDE concentrations ranged from 15.35 ng/g to 45.34 ng/g. Mean 4,4’-DDD concentrations ranged from 28.16 ng/g to 63.25 ng/g and mean 4,4’-DDT concentrations ranged from 28.64 ng/g to 96.21 ng/g. Notably, 4,4’-DDE and 4,4’-DDT were accumulated to the highest concentration in the exposure tank within the highest nominal concentration (63.17 ng/mL). This was however not the case with 4,4’-DDD which was accumulated to the highest concentration in the exposure tank with the second highest nominal concentration (31.58 ng/mL).

Table 5. 3: Levels of accumulation in S. zambezensis muscle tissue and concentrations in water samples of the 4,4’-DDT degradation products (mean ± SD) and range after an acute laboratory exposure. Exposure 4,4’-DDE 4,4’-DDD 4,4’-DDT 4,4’-DDE (ng/g) 4,4’-DDD (ng/g) 4,4’-DDT (ng/g) Conc. (ng/mL) (ng/mL) (ng/mL) (ng/mL) 15.34 ± 1.49 28.16 ± 3.92 28.64 ± 2.57 0.43 ± 0.08 3.98 ± 0.21 0.75 ± 0.07 1.97 (11.18 - 15.52) (19.55 - 40.20) (22.39 - 36.27) (0.34 - 1.24) (2.95 - 4.99) (3.40 - 3.46) 35.32 ± 8.75 42.50 ± 4.83 39.90 ± 7.45 5.91 ± 0.63 7.27 ± 0.55 7.74 ± 0.40 3.95 (11.26 - 80.31) (25.28 - 67.59) (0.98 - 67.94) (3.33 - 9.30) (4.47 - 9.87) (5.50 - 9.29)

20.88 ± 2.18 43.73 ± 7.00 47.72 ± 7.46 2.51 ± 0.16 3.99 ± 0.14 5.13 ± 0.16 7.90

(13.07 - 32.52) (17.35 - 74.86) (22.00 - 87.20) Water (1.71 - 3.32) (3.51 - 4.90) (4.60 - 6.06) Tissue 29.43 ± 2.76 48.87 ± 2.62 54.12 ± 9.06 2.04 ± 0.05 3.80 ± 0.30 5.33 - 0.25 15.79 (19.59 - 41.18) (39.55 - 61.69) (1.65 - 87.48) (1.92 - 2.26) (3.26 - 6.81) (4.67 - 7.43) 40.70 ± 4.90 63.25 ± 8.50 75.32 ± 8.27 10.50 ± 2.08 11.19 ± 1.69 9.59 ± 0.71 31.58 (20.89 - 62.78) (44.76 - 114.53) (52.26 - 124.75) (4.12 - 25.39) (5.83 - 23.42) (6.72 - 10.77) 45.34 ± 5.21 60.29 ± 2.49 96.21 ± 6.41 2.85 ± 0.24 4.12 ± 0.21 7.38 ± 0.99 63.17 (33.72 - 72.83) (54.63 - 75.88) (72.57 - 128.83) (1.97 - 4.31) (3.38 - 5.55) (4.84 - 16.34) It is clear that S. zambezensis accumulates 4,4’-DDT and its degradation products from its immediate environment, in this case the exposure tank water. The results listed in Table 5.4, show that from a nominal concentration of 1.97 ng/mL (lowest) to 63.17 ng/mL (highest) a 4,4’-DDT equivalents of 76.92 ng/g and 213.53 ng/g respectively were accumulated after the acute exposure period. Should the exposure period be increased, even larger accumulated amounts of 4,4’-DDT equivalents are expected.

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Table 5. 4: Levels of accumulation of the 4,4’-DDT degradation products in S. zambezensis muscle tissue and extracted from the exposure tank water as well as percentage 4,4’-DDT accumulated in S. zambezensis. The 4,4’-DDT used was of 98.7% purity purchased from Sigma-Aldrich PESTANAL® as an Analytical Standard. Exposure Tanks Average accumulated by S. zambezensis (ng/g) Average extracted from water after 96-hrs (ng/mL) Conc. Total DDT 4,4’-DDT % initial 4,4’-DDT % initial 4,4'-DDE 4,4’-DDD 4,4’-DDT 4,4'-DDE 4,4’-DDD 4,4’-DDT (ng/mL) in tank (μg) Equivalents dose Equivalents dose 1.97 98.70 15.34 28.16 28.64 76.92 77.93 0.43 3.98 0.97 5.86 5.94 3.95 197.40 35.32 42.50 39.90 126.34 64.00 5.91 7.27 7.74 22.39 11.34 7.90 394.80 20.88 43.73 47.72 119.43 30.25 2.51 3.99 5.13 12.34 3.13 15.79 789.60 29.43 48.87 54.12 141.05 17.86 2.04 3.80 5.33 11.81 1.50 31.58 1579.20 40.70 63.25 75.32 190.75 12.08 10.50 11.19 9.59 33.69 2.13 63.17 3158.40 45.34 60.29 96.21 213.53 6.76 2.85 4.12 7.38 15.11 0.48 The trend seen is that the accumulated concentrations of each degradation product seems to increase as the nominal concentration increases from the lowest nominal concentration (1.97 ng/mL) to the highest nominal concentration (61.17 ng/mL) (Figure 5.1). The accumulated concentrations of each degradation product increases as the nominal concentration increases, however the total percentage accumulated (Table 5.4) of the initial dose of 4,4’-DDT decreases as the nominal concentrations increases.

5.2.4 Levels of 4,4’-DDT in water samples In the current study, the three degradation products found were 4,4’-DDE (0.34 ng/mL to 25.39 ng/mL), 4,4’-DDD (2.95 ng/mL to 23.42 ng/mL) and 4,4’-DDT (3.40 µg/L to 16.34 µg/L), and unlike in the fish tissue, 4,4’-DDD is found at the highest concentrations (Table 5.3). The general trend sees the constant presence of all three degradation products with the highest concentrations of the 4,4’-DDE and 4,4’-DDD being found in the second highest nominal concentration. The highest concentration of the parent compound 4,4’-DDT was however found in the highest nominal concentration.

When looking at the concentrations at 24-hour intervals, there is no clear trend for the various concentrations in exposure tanks, but the duplicate tanks showed the same trend in each concentration, giving confidence that the extraction and analysis of the samples were consistent throughout. The presence of a low standard deviation (SD) is also an indication of the consistency.

Each exposure tank, regardless of the nominal concentration added, shows the presence of all 3 degradation products, with little variation between the 24-hour

60 | P a g e sample periods. The general trend was that the concentrations of each metabolite seemed to decrease over the 96-hour period. This decrease over time may be linked to increased degradation and adsorbtion to colloidal particulates within exposure tank water (Taha and Mobasser, 2014).

At the highest exposure concentration, it is notable that although all three degradation products are present, the concentrations are less than half those found in the second highest concentration. The duplicate tanks at the highest concentration contained large amounts of secreted mucus from the fish. This mucus was removed from the water samples before extraction and analysis and this may be the reason why these degradation products do not occur at these high concentrations.

The parent compound 4,4’-DDT occurs at all intervals in all tanks and indicates the “recent” nature of this exposure.

5.2.5 Oxidative stress biomarkers The resulting biomarker responses in the liver and gill tissue are reported in Figure 5.1A-E, with significant differences (p≤0.05) being indicated by common superscripts (capital or small letters of the alphabet). Unfortunately MDA analysis could only be performed in the liver tissue as the gills were too small to allow for the analysis of MDA as well as the other oxidative stress biomarkers. The liver and the gills were used because the liver is the detoxifying organ of the organism and the gills are in direct contact with the aquatic environment and may show more advanced changes during an acute exposure (Rola et al., 2012).

The SOD analysis resulted in varied responses between tissues. The SOD response in the liver tissue ranged from a mean of 0.93 ng SOD/mg protein to 1.57 ng SOD/mg protein, while in the gills the response was much higher and ranged from a mean of 2.05 ng SOD/mg protein to 4.06 ng SOD/mg protein. There were no significant differences (p≤0.05) noted within the liver at the varying concentrations, but within the gills significantly higher concentrations of SOD were found at the 7.90 ng/mL and 31.50 ng/mL exposure concentration when compared to the control. A significantly higher SOD response was also noted in the gills at the 31.58 ng/mL exposure concentration when compared with the 15.79 ng/mL exposure concentration (Figure 5.1A).

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The MDA (nmol/mg protein) concentrations ranged from a mean of 0.11 nmol/mg protein to 0.66 nmol/mg protein and a significant difference (p≤0.05) is noted between the 3.95 ng/mL and 7.90 ng/mL exposure concentrations, with a significantly lower resonse in the higher exposure concentration (Figure 5.1E).

Figure 5. 1: Biomarkers of Oxidative stress within the liver and the gills of Synodontis zambezensis after an acute exposure to 4,4’-DDT. Superoxide dismutase (Figure 1A), PC (Figure 1B), CAT (Figure 1C), GSH (Figure 1D) and MDA (Figure 1E). Significant differences (p≤0.05) between tissues within the same biomarker are indicated by the presence of capital letter suberscripts. Significant diffierences (p≤0.05) within the same tissue (liver or gills) of the same biomarker are indicated by common lower case letter suberscripts. Protein carbonyl responses were similar when comparing the gills and the liver and ranged from a mean of 185.07 nmol carbonyl/mg protein to 255.34 nmol carbonyl/mg

62 | P a g e protein and from 213.79 nmol carbonyl/mg protein to 272.84 nmol carbonyl/mg protein respectively. There was a significant difference (p≤0.05) noted only in the liver tissue with a significantly lower PC concentration in the 1.97 ng/mL exposure concentration when compared with the control (Figure 5.1B).

The CAT concentrations ranged from 26.17 µmol H2O2/min/mg protein to 57.86 µmol

H2O2/min/mg protein in the gills and from 150.62 µmol H2O2/min/mg protein to

274.01 µmol H2O2/min/mg protein in the liver. The response of CAT was clearly greater within the liver tissue, however there was no significant difference (p≤0.05) found between exposure concentrations in the liver tissue (Figure 5.1C). In the gill tissue significantly lower CAT responses were noted at the 7.90 ng/mL, 31.58 ng/mL and 63.17 ng/mL exposure concentrations when compared to the control. A significantly higher response in CAT was found at the 1.97 ng/mL exposure concentration when compared with the response seen at the 7.90 ng/mL and 31.58 ng/mL exposure concentrations. The final significant difference seen in CAT was a higher response at the 3.95 ng/mL exposure concentration when compared to the 7.90 ng/mL and 31.58 ng/mL exposure concentrations.

Reduced glutathione concentrations within the gill tissue showed a significantly higher response at the 63.17 ng/mL exposure concentration when compared to all other exposure concentrations and the control (Figure 5.1D). In the liver tissue a significantly higher response was seen at the 1.97 ng/mL and 3.95 ng/mL exposure concentrations when compared to the 15.79 ng/mL and the 31.58 ng/mL exposure concentrations. The GSH concentrations in the gills ranged from a mean of 1.95 µg/mg protein to 4.67 µg/mg protein. The mean range of GSH in the liver tissue was from 1.26 µg/mg protein to 2.26 µg/mg protein.

5.2.6 Relationship between biomarkers and bio-accumulation In order to relate the biomarkers of oxidative stress to 4,4’-DDT and its degradation products, an RDA triplot was constructed (Figure 5.2). An RDA can be considered a constrained version of a principal component analysis (PCA), meaning that the responses (oxidative stress biomarkers) are constrained by the 4,4’-DDT and the degradation products (Quinn and Keough, 2002). This constrained ordination gives an indication of the effect of 4,4’-DDT and its degradation products on the spread of the oxidative stress biomarkers within the two focus tissues. The angles between all vectors reflect their (linear) correlation. The correlation between the responses

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(oxidative stress biomarkers), the contsraining variables (4,4’-DDT and its degradation products) or any combination thereof, is equal to the cosine of the angle between vectors (arrows) (Quinn and Keough, 2002). For example, should the angle between two vectors be at an angle equal to 90°, the two vectors are uncorrelated since cos(90) = 0. However, should the angle between two vetors equal 20°, these two vectors will have a strong, positive correlation since cos(20) = 0.94 (Quinn and Keough, 2002). The length of the arrows is also an indication of the strength or influence of the vector, with longer arrows having a greater influence (Quinn and Keough, 2002).

The ordination in Figure 5.2 explains 94.41% of the variance within the data. The first axis (function 1 – Horizontal axis) explains 35.77% of the variance within the data. The second axis (function 2 – Vertical axis) explains 58.64% of the variance within the data and a clear separation is seen vertically between SOD in the gills and liver, GSH in the gills, and CAT in the liver tissue when compared with the other biomarkers of oxidative stress. The arrow for 4,4’-DDT is much longer than that of its degradation products and this indicates its increased influence on the resulting biomarker excitation.

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GSH (L) 0.4

CAT (G)

PC (L) PC (G)

CAT (L)

MDA (L) SOD (G)

SOD (L) GSH (G)

4,4'-DDE 4,4'-DDD

4,4'-DDT -1.0 -0.6 0.8

Figure 5. 2: RDA triplot showing resulting oxidative stress biomarker responses in Synodontis zambezensis following an acute exposure to varying concentrations of 4,4’-DDT. The resulting oxidative stress biomarkers are constrained in this ordination by 4,4’-DDT and its degradation products, 4,4’-DDE and 4,4’-DDD. The angle between SOD in the gills and liver as well as GSH in the gills indicates a strong correlation of these biomarkers with 4,4’-DDT (≤90º). The same can be seen with 4,4’-DDD and 4,4’-DDE with the above mentioned biomarkers. The angles between 4,4’-DDT and its degradation products and MDA in the liver are ≥90 ˚, this indicates a weak correlation between the biomarker response and 4,4’-DDT and its degradation product. This can be seen with PC and GSH in the liver, as well as CAT and PC in the gills.

Table 5. 5: Tabulated correlation coefficients (factor or component loadingds) between the resulting oxidative stress biomarkers (SOD,MDA, PC, CAT and GSH) in both the liver and gill tissue of Synodontis zambezensis in relation to 4,4’-DDT and its degradation products 4,4’-DDE and 4,4’-DDT. Liver Gills 4,4'-DDE 4,4'-DDD 4,4'-DDT 4,4'-DDE 4,4'-DDD 4,4'-DDT SOD 0.57 0.50 0.49 0.15 0.18 0.28 MDA -0.21 -0.08 0.11 NA NA NA PC -0.24 0.15 0.02 0.09 0.11 0.02 CAT 0.34 0.24 0.17 0.10 -0.12 -0.17 GSH -0.11 -0.17 -0.33 0.46 0.27 0.43

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The correlation coefficients glisted in Table 5.5, give a clearer indication of the correlations between the different oxidative stress biomarkers in the liver and gill tissues, and 4,4’-DDT and its degradation products seen in the RDA triplot. Correlation coefficients are values which range between +1.0 and -1.0, and give an indication of the relationship between two sets of variables (Quinn and Keough, 2002), for example SOD in the gills and 4,4’-DDT. In this example there is a positive correlation indicated by a correlation coefficient of 0.28, and this is corroborated by the ≤90º angle seen between these two arrows in the RDA triplot (Figure 5.2). The same justification can be made by comparing the correlation coefficients listed in Table 5 and the angles between the various arrows in the RDA triplot in Figure 5.2.

5.3 Discussion 5.3.1 Water quality The TDS within aquatic systems is contributed to by both inorganic and organic dissolved salts (Thirumalini and Joseph, 2009), therefore higher dissolved salts leads to higher TDS readings. The EC of a water sample can be correlated with TDS but gives slightly different information, as it relates to the electrical nature of the dissolved salts within the water (Atekwana et al., 2004). Both the TDS and the EC readings increased as the exposure concentrations of 4,4’-DDT increased in the relevant exposure tanks. The addition of contaminants from the 4,4’-DDT may have contributed to the increase in EC (Thirumalini and Joseph, 2009), but the increases in TDS are much more significant across the exposure tanks. Since TDS readings are affected by both organic and inorganic salts, the increase in mucus secreted by the fish in the tanks containing higher 4,4’-DDT concentrations may have affected the TDS (Church et al., 2009). Fish mucus is composed mainly of large macromolecules which form a gel in water (Easy and Ross, 2009), these macromolecules may interact with dissolved ions and affect the total TDS readings (Easy and Ross, 2009).

5.3.2 Levels of 4,4’-DDT in Synodontis zambezensis The decrease in total percentage accumulated DDTs (Table 5.4) as the nominal concentrations increases can be attributed to various factors that may include increased volatisation through photodegradation, adsorption to the tank surfaces and sedimentation resulting in less bio-availability of the metabolites (Foght et al., 2001). Volatisation of 4,4’-DDT occurs almost immediately after it enters the aquatic

66 | P a g e environment and the reported half-life for 4,4’-DDT is approximately 56 days in lakes and 28 days in riverine systems (Bedos et al., 2002; USEPA, 1989).

The only routes of accumulation present in this study were directly through either the skin or respiration through the gills. It is, therefore, no surprise that the bio- accumulation of the parent compounds (4,4’-DDT) is much higher than that of the more degraded forms (DDE and DDD) (Yohannes et al., 2014) in all of the exposure tanks. Higher concentrations of 4,4’-DDE and 4,4’-DDD are associated with biomagnification through the food chain since they are a result of the physiological breakdown of 4,4’-DDT through animal metabolic processes (Rognerud et al., 2002). The general pattern of accumulation seen in the wild is that 4,4’-DDE is the highest, followed by 4,4’-DDD and then 4,4’-DDT (Bedos et al., 2002). There was some metabolic breakdown in this study, but since the exposure was acute, the exposed fish did not metabilise the 4,4’-DDT to the same extent as they would have over a longer period of time. A study conducted by Macek and Korn (1970) revealed the importance that the food chain plays in the accumulation of 4,4’-DDT within fish. In the study, the results show that up to 35.5% of the 4,4’-DDT available to Salvelinus fontinalis (brook trout) was accumulated when it was exposed through its diet. On the other hand, only 3.55% of the total 4,4’-DDT available was accumulated when exposure took place only through the water (Kidd et al., 2001).

An observation made during the exposure period was that the higher concentrations caused the fish to secrete large quantities of mucus into the water. This mucus can be secreted as a response to pathogenic agents or xenobiotics in their environment, and the quantity and nature of the mucus they secret can vary in relation to the changes in the environment (Easy and Ross, 2009). The presence of these high volumes of mucus may have increased the surface area for adhesion of 4,4’-DDT and its degradation products, making the bio-availability lower in the tanks. This means that even though the nominal concentrations were highest in the tanks, the ability for accumulation to occur was decreased (Foght et al., 2001). Mucus has also been shown to serve a protective function in fish, creating a barrier between the fish and the external environment (Church et al., 2009), this may in turn decrease the ability for the fish to bio-accumulate toxicants. As seen in Table 4, in the lowest nominal concentration, 73.08% of the 4,4’-DDT added to the tank was accumulated by S. zambezensis, and this percentage decreases as the nominal concentration

67 | P a g e increases. In the highest nominal concentration, only 6.39% of the initial 4,’4-DDT was accumulated by S. zambezensis, another indication that although the overall accumulation increased as the nominal concentration increased, the percentage accumulated still decreases.

Synodontis sp. is known for its high tissue fat content which has been previously linked to their higher than normal buoyancy and results in their ability to achieve their well documented upside-down swimming behaviour (Lalèyè et al., 2006). Since 4,4’- DDT is lipophilic and accumulates in fatty tissues (Gaspar et al., 2015) it is expected that fish with higher fat contents will accumulate higher concentrations of 4,4’-DDT (Kidd et al., 2001). Unfortunately, the full extent of S. zambezensis accumulation is not clear from the results of this study since the exposure period was acute and because the addition of bio-magnification through the food chain is not illustrated here as it was not within the scope of the study. It is clear from this study, however, that S. zambezensis does accumulate 4,4’-DDT and its degradation products at significant concentrations, even only after a single exposure and during a 96-hour period. The accumulated concentrations found during this study fall within the environmentally relevant concentrations found in the same species in the lower Phongolo River and floodplain (Smit et al., 2016). In the study by Smit et al. (2016) 4,4’-DDT concentrations ranged from 0.07 ng/g to 802.17 ng/g and 4,4’-DDE concentrations ranged from 0.09 ng/g to 1208.18 ng/g. The lowest degradation product found was 4,4’-DDD and ranged from 0.07 ng/g and 567.07 ng/g. Although these concentrations were much higher than those found in the current study, the concentrations are within the environmental range of accumulation. Conversely, 4,4’- DDE in the study by Smit et al. (2016) is much higher and is a clear reflection of the effect of the food chain on accumulation in this fish species (Smit et al., 2016).

5.3.3 Levels of 4,4’-DDT in water samples A microcosm study done by the USEPA (USEPA, 1979) illustrates the fate of 4,4’- DDT in the aquatic environment after 30-40 days. The 4,4’-DDT concentrations present in the water column had decreased below detectable limits and that 90% of the initial 4,4’-DDT added was not present in the fish, invertebrates, algae or sediment within the pond and was therefore presumed to have volatised during the course of the experiment. During the first 30 days, 4,4’-DDT was the main

68 | P a g e degradation product present. The results of this study reflect those found in the microcosm study (USEPA, 1979).

The formation of 4,4’-DDE takes place through photochemical reactions, which require the presence of sunlight (Leaños-Castañeda et al., 2007) as well as through bacterial dehydrochlorination (Leaños-Castañeda et al., 2007) and animal dehydrochlorination (Kitamura et al., 2002). The breakdown of 4,4’-DDT after addition to the exposure tanks may have been affected by the day/night cycle present within the environmental room as well as bacteria present within the borehole water and on the skin of S. zambezensis. For this reason, 4,4’-DDE may not have been very high within the fish tissue, but the degradation processes within the water increased the presence of 4,4’-DDE. The presence/absence of some degradation products and not others is unclear, but the previously mentioned adhesion and adsorption properties of 4,4’-DDT and it’s degradation products may be responsible for this (Pontolillo and Eganhouse, 2001). The general trend observed was that the concentrations of each metabolite decreases over the 96-hour period, and may be attributed to the degradation of 4,4’-DDT in aquatic systems into its degradation products (Taha and Mobasser, 2014). The mucal secretions could also have contained higher concentrations of the 4,4’-DDE and 4,4’-DDD than the fish tissue because they were secreted by the fish and could have been a mechanism to release 4,4’-DDT and any of its degradation products from their bodies (Church et al., 2009). Since these secretions were not tested separately, in this case, it can not be confirmed whether or not this is a valid assumption.

5.3.4 Oxidative stress biomarkers and the link to bio-accumulation The formation of ROS within both the liver and the gills of S zambezensis is a physiological response to the exposure to 4,4’-DDT (Abdollahi et al., 2004). The formation of antioxidant species in response to exposure to various xenobiotics has been widely documented in a variety of aquatic organisms (Ferreira et al., 2005; Van der Oost et al., 2003).

The constant presence of PC within the liver tissue of S. zambezensis from the control throughout the exposure concentrations is an indication that there may not have been high levels of protein breakdown within the organism (Parvez and Raisuddin, 2005). This does not necessarily mean that there was no potential for protein damage, but the short length of the exposure period may have reduced the

69 | P a g e physiological adaptation of PC production from taking full effect (Connell et al., 2016). There were no significant differences (p≤0.05) noted in PC concentrations in the gill tissue of S. zambezensis, which is surprising considering their proximity to the 4,4’-DDT in the exposure tank (Rola et al., 2012). The concentrations of PC reported in a study by Smit et al. (2016) are similar to those found during this study. Synodontis zambezensis in the study by Smit et al. (2016) had much higher PC concentrations than the other three fish species studied. This may be an indication that S. zambezensis has naturally higher PC concentrations than those found in other fish species. The RDA ordination shows a negative correlation between 4,4’- DDT and its degradation products and PC activity indicated by the ≥90 ˚ angle.

Oxidative stress is clearly seen to occur in the liver tissue of S. zambezensis through the presence of higher concentrations of MDA from the control to the exposure concentrations that follow. Lipid peroxidation is often a precursor to protein damage (Traverso et al., 2004) and this could be another indication of the effect that the length of the exposure has on the formation of particular ROSs and antioxidants (Connell et al., 2016). Since MDA is a precursor to PC, the action of MDA within the liver may be preventing further damage from occurring within the proteins further down the oxidative cycle (Traverso et al., 2004).

Superoxide dismutase and CAT work together in an effort to balance out the oxidation caused by ROSs and are often referred to as the first line of defence in the presence of oxidative stress (Van der Oost et al., 2003), and this can clearly be seen in the gill tissue of S. zambezensis. As the SOD activity increases from the control to the exposure concentrations, CAT activity decreases. This balancing effect is also clear in the RDA with increased length of SOD and CAT arrows and direct correlation (<90˚ angle) with 4,4’-DDT and its degradation products. The significantly higher concentration of CAT in the gills at the lower concentration of 4,4’-DDT indicates an initial adaptive response by the fish in this concentration (Abdollahi et al., 2004).

Reduced glutathione plays an important role in both direct and enzymatic neutralization of ROSs (Giustarini et al., 2013). Its activity can be seen in both the liver and gills of S. zambezensis, but its effect is more apparent in the gills. At the highest exposure concentration, the GSH activity is significantly higher (p≤0.05) than the control and all other (lower) exposure concentrations. The RDA ordination also

70 | P a g e shows the importance of GSH activity in the gill tissue with its close proximity (<90˚ angle) to 4,4’-DDT and its degradation products. A more conclusive way of measuring GSH activity within an organism is to compare the production of GSH to its oxidised form GSSG (Giustarini et al., 2013), and could be considered in future studies.

5.4 Conclusions and recommendations This is the first laboratory-based study to report on accumulated concentrations of 4,4’-DDT, its degradation products and the resulting oxidative stress biomarkers within Synodontis zambezensis. Based on the results it is clear that even an acute exposure to relatively low concentrations of 4,4’-DDT, can result in some accumulation without the contribution of biomagnification through the food chain. Although chronic exposure periods are considered to give a more accurate indication of the physiological responses in aquatic organisms, an acute exposure such as this has proven that oxidative stress may still result after a short exposure to xenobiotics. These results reflect those which may occur during a spill event or during spray contamination of nearby water sources. The exact mechanisms of metabolism and the enzymatic systems involved in the breakdown and storage of 4,4’-DDT and its degradation products, in this fish species, have not been previously documented, and so it is recommended that further studies e.g. chronic exposure of this species to 4,4’-DDT be completed. Chronic exposure experiments may also allow for the species specific Bioconcentration Factor (BCF) for 4,4’-DDT to be calculated (USEPA, 1996). Since the uptake phase in this study was not 28 days and the ratio between uptake and depuration was not noted, it is not possible to calculate the species specific BCF from the current data. It is further recommended that the battery of oxidative stress biomarkers be increased to include the ratio calculation of GSSG to GSH, specifically within the gill tissue. The inclusion of gill histology studies is also recommended to determine whether structural changes have taken place within the gills since they are in direct contact with the toxicant (Rola et al., 2012). Identification of possible changes may enhance the understanding of oxidative stress within the gill tissue.

The interactions that have taken place between the 4,4’-DDT, the exposure borehole water and the conditions within the environmental room are unclear and further

71 | P a g e investigation into the partitioning, volatilisation and adsorption of 4,4’-DDT needs to be investigated.

5.6 Acknowledgements This study was supported through funding by the National Research Foundation (NRF – UID: 102440), Hokkaido University, Sapporo Japan (UID 92424 Japan/ SA Bi-lateral Programme) and the University of Johannesburg (Global Excellence Stature Scholarship). Mr Harris and Dr Meyer from Shimadzu South Africa are acknowledged for the use of their equipment.

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Chapter 6: Biomarkers of exposure and energetics

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6. Introduction Biomarkers reflect negative physiological changes in response to exposure to xenobiotics (Van der Oost et al., 2003). Biomarkers are normal enzymatic responses present within organisms, changes to which, can indicate a negative influence caused by the immediate external environment (Cormier and Daniel, 1994). Biomarkers responses are coping mechanisms found in organisms in response to the effects of stressors in the environment (Cormier and Daniel, 1994) and can be measured in a variety of fluids and tissues, from blood serum to muscle tissue, depending on the biomarker response being measured (Bucheli and Fent, 1995). Biomarkers of effect, specifically those relating to oxidative stress caused by an imbalance in reactive oxygen species (ROSs) and their respective antioxidants (Almroth et al., 2008), have been discussed in detail in Chapter 5. Biomarkers of exposure form the focus of this chapter and they are acetylcholinesterase (AChE), cytochrome P450 (CYP450) and cellular energy allocation (CEA).

Acetylcholinesterase has been studied extensively in a variety of organisms (Sturm et al., 1999), resulting in its effective and well established use as a biomarker of exposure. Acetylcholinesterase, which functions in neuromuscular junctions and cholinergic synapses, can be irreversibly inhibited by the presence of OCPs, a group that includes 4,4’-DDT (Sturm et al., 1999). The physiological role of AChE is in breaking down ACh at the end of a muscle contraction (Solé et al., 2010). Inhibition of AChE can lead to an increase in muscular spasms and eventual death of the affected organism (Weiss and Gakstatter, 1964). The inhibition of AChE is an indicator of exposure to pesticides and carbamates, and these levels can be measured in a variety of organs including the muscle tissue, brain tissue and liver tissue (Leibel, 1988; Lundin, 1962).

Cytochrome P450 (CYP450) is a biomarker, most widely used to indicate detrimental changes caused through xenobiotic exposure (Bucheli and Fent, 1995). Polar organic pollutants that enter organisms are extensively bio-transformed before they can be excreted and this bio-transformation takes place in two stages (Andersson and Förlin, 1992). The CYP450 super family consists of monooxygenases, which detoxify lipophilic chemicals through the introduction of oxygen (Ioannides and Parke, 1990) this is known as phase 1 bio-transformation (Andersson and Förlin, 1992). The resulting oxidised pollutant is easily excreted through the action of

74 | P a g e endogenous molecules during phase 2 bio-transformation (Andersson and Förlin, 1992). Species sensitivity can be directly linked to the function and regulation of CYP450 proteins within an individual organism (Morrison et al., 1995), meaning that more susceptible species have fewer functioning CYP450 genes (Morrison et al., 1995). Cytochrome P450 is found in animals, plants and microorganisms, and is predominantly found in the liver tissue in mammals (Porter and Coon, 1991). Metabolism of xenobiotics is not restricted to the liver (Andersson and Förlin, 1992) and this is evident in fish since xenobotic metabolism has been noted to take place in other organs and parts of the body, including the kidneys and gills (Varanasi et al., 1982).

Toxic stress may result in an imbalance in the metabolism of affected organisms (De Coen et al., 1995), resulting in measurable changes in the available energy (Ea) and the consumed energy (Ec) (De Coen and Janssen, 1997). The Ea and Ec can be combined to give a final indication of the total cellular energy allocation (CEA) within an organism (De Coen and Janssen, 1997). Cellular energy allocation is used as a biomarker and allows for changes in physiological energetics to be monitored in relation to increasing xenobiotic exposure (Verslycke et al., 2004). Energy consumption is a measurement of the electron transport system (ETS) in the mitochondria of muscle tissue (De Coen and Janssen, 1997). The ETS functions by passing electrons along a series of membrane proteins resulting in the generation of adenosine triphosphate (ATP), the energy currency in cells (Cammen et al., 1990). Measuring ETS activity gives an indication of the estimated cellular respiration and therefore metabolism which has taken place in an organism (Cammen et al., 1990). The Ea is determined through the separate determination of glucose (carbohydrates), lipids and proteins present in the muscle tissue. Together these concentrations represent the total amount of energy available to the organism (De Coen and Janssen, 1997). The final determination of CEA is completed by subtracting the Ea from the Ec (De Coen et al., 1995; De Coen and Janssen, 1997).

The aim of this section of the study was to investigate whether an acute exposure to varying concentrations of 4,4’-DDT would result in any changes in the above mentioned exposure biomarkers in Synodontis zambezensis. The exposure concentrations and the methods used during the acute exposure experiment have been outlined in Chapter 5. These concentrations are used for comparisons in this

75 | P a g e chapter as well, since the fish tissue used was analysed for both biomarkers of effect and exposure. The aim was completed through the use of the methods and procedures outlined in the materials and methods.

6.1 Materials and methods 6.1.1 Fish tissue collection Liver tissue was removed and its mass was recorded with an electronic scale (Sartorius Basic BA210S). Muscle tissue was removed from the lateral side of each fish and weighed (Sartorius Basic BA210S). All tissue was placed into separate falcon tubes containing Henriksson’s stabilising buffer (Henriksson et al., 1986). Muscle tissue was used for biomarker analysis of AChE and CEA and liver tissue was used for biomarker analysis of AChE and CYP450. All samples were frozen at - 80˚C until further analysis.

6.1.2 Biomarker protocols Protein content was determined for each biomarker in both tissues (Bradford, 1976). Absorbance was measured at 630 nm and bovine serum albumin (BSA) was used as a reference protein standard for determination of protein content. All biomarker activities are expressed per milligram protein to allow for standardisation.

Biomarkers of exposure Acetylcholinesterase Methods used of the analysis of AChE were adapted from Ellman et al. (1961). Acetylcholinesterase was measured both in the liver and muscle tissue. The normal protocol followed (Ellman et al., 1961), refers to the analysis of AChE in the liver tissue only, but since AChE is an important enzyme involved in the functioning of the neuromuscular junction (Sturm et al., 1999), the analysis of this biomarker in the muscle tissue seemed appropriate for this study. Liver and muscle tissue were homogenised in separate Eppendorf tubes. A mass of 0.2 g of each tissue was homogenised in Tris/sucrose buffer (pH 7.4). The homogenate was centrifuged (High Speed Refrigerated Centrifuge: NovaFuge B113-21R, Senova Biotech Co.,Ltd) at 9 500 g., 4 ˚C for 10 mins and the supernatant was used for AChE and protein content analysis. Only 24 wells of the 96-well microtiter plate were used for analysis since the reaction is light sensitive and kinetic. Therefore, only seven samples in triplicate were analysed per plate. Two hundred and ten microliters of potassium phosphate buffer (pH 7.4), 10 µL 2-(acetylsulfanyl)-N,N,N-trimethylethanaminium iodide (s-

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Acetylthiocholine iodide, 98% Sigma-Aldrich) and 10 µL 2,2′-dithiodibenzoic acid (Ellman’s reagent, ≥95%, Sigma-Aldrich) was added to each of the 24 wells. The plate was tapped gently to ensure mixing, covered and incubated for 5 mins at 37˚C. The blank used for analysis was 5 µL Tris/Sucrose buffer, added to the first three wells. To the remaining wells 5 µL of each sample was added in triplicate. The plate was analysed at 405 nm for 6 mins at 1 min intervals using an automated microplate reader (Elx800-Universal Microplate Reader, BioTek® Instruments, Inc.). The change of absorbance over time was measured since this is a kinetic reaction.

Cytochrome P450 The analysis of CYP450 was completed using a DetectX® P450 Fluorescent Activity kit (Arbor Assays®). Zero point one grams of liver tissue was homogenised in 1 mL of GHB (pH 7.4). Microtiter plates and solutions were provided in the kit and the manufacturer’s instructions were followed for completion of the analysis. Ninety five microliters of the blank, standards and each sample were added to the microtiter plates in triplicate, covered and incubated for 15 mins at 37˚C. Following the incubation period 5 µL (2R,3R,4R,5R)-5-(6-aminopurin-9-yl)-3-hydroxy-4- phosphonooxyoxolan-2-yl methoxy-hydroxyphosphoryl (2R,3S,4R,5R)-5-(3- carbamoyl-4H-pyridin-1-yl)-3,4-dihydroxyoxolan-2-yl]methyl hydrogen phosphate (NADPH, Reduced nicotinamide-adenine dinucleotide phosphate) was added to each well and further incubated at 37˚C for 30 mins. A stop solution (5 µL) and Formaldehyde Detection Reagent DetectX® (25 µL) were added and a final incubation period at 37˚C for 30 mins was completed. Each plate was analysed at 510 nm with excitation at 450 nm, for the determination of florescence, using a Multi- Detection microplate reader (FLx800 Fluorescence Microplate Reader, BioTek® Instruments, Inc.).

Biomarkers of energetics Cellular energy allocation The methods used for the analysis of CEA were carried out according to the protocols outlined in De Coen and Janssen (1997, 2003). Muscle tissue is used for the analysis of total CEA, which is split into two separate protocols, Energy availability (Ea) and Energy consumption (Ec). Energy availability is further split up into the separate analysis of protein content (Bradford, 1976), glucose content and lipid content. The energy consumed is analysed through the electron transport chain

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(ETS). Muscle tissue was weighed out separately for Ea and for Ec. A mass of 0.1 g of muscle tissue was homogenised in 400 µL deionised water for Ea and 0.1 g of muscle tissue was homogenised in 400 µL of ETS buffer (pH 8.5) for Ec.

Energy availability Since the energy available in any living organism is split into three different stored types of energy (protein, lipids and carbohydrates), these three components need to be analysed separately in order to determine the overall energy available in an organism (De Coen and Janssen, 1997).

The determination of protein content followed the methods outlined in Bradford (1976). Five microliters of each homogenised sample was added in triplicate to a microtiter plate, followed by the addition of 245 µL Bradford reagent (Bradford, 1976). The reference standard used for this protocol is bovine serum albumin (BSA) and the blank is deionised water. Each microtiter plate was analysed at 595 nm using an automated microplate reader (Elx800-Universal Microplate Reader, BioTek® Instruments, Inc.).

The carbohydrate or glucose content of the muscle tissue was determined using a glucose content test kit (GOD-PAP 1 448 668, Roche). All solutions used were supplied in the test kit and the manufacturer’s instructions were followed for complete analysis. A volume of 2.5 µL of the homogenate of each sample was pipetted in triplicate into the wells of a microtiter plate. To each sample 247.5 µL of the supplied glucose standard (C FAS 759 350, Roche) was added. The blank used for this protocol was deionised water. The microtiter plates were incubated at room temperature for 30 mins and absorbencies were analysed at 560 nm using an automated plate reader (Elx800-Universal Microplate Reader, BioTek® Instruments, Inc.).

The extraction and purification of total lipid content was completed (Bligh and Dyer, 1959) following the addition of 250 µL homogenate to 500 µL chloroform (≥99.9%, Sigma-Aldrich). This mixture was added together in a separate Eppendorf tube and vortexed (Finevortex, FINEPCR®). To the mixture, 500 µL methanol (≥99.9%, Sigma-Aldrich) and 250 µL deionised water were added and vortexed again. The resulting mixture was centrifuged (High Speed Refrigerated Centrifuge: NovaFuge B113-21R, Senova Biotech Co.,Ltd) at 4˚C, 2 940 g. for 10 mins. Following

78 | P a g e centrifugation, two distinct layers were visible within each sample with the organic phase being present in the bottom layer. One hundred microliters of the organic phase was pipetted into separate glass test tubes. To each glass test tube, 500 µL sulphuric acid (H2SO4, ≥99.9%, Sigma-Aldrich) was added and carefully vortexed. Each test tube was covered in a small tin foil lid and placed into an oven to incubate at 200˚C for 15 mins. The test tubes were removed from the oven, allowed to reach room temperature and 1 mL of deionised water was carefully pipetted into each test tube. For the analysis of these samples polyethylene microtiter plates (Thermo Scientific) where used to combat the corrosive effects of the sulphuric acid. The blank used for this protocol was 1,2,3-propanetriyl trihexadecanoate (Tripalmitin, 99%, Sigma-Aldrich) and a volume of 245 µL was added in triplicate to each plate. The same volume of each sample was added in triplicate to the microtiter plate and the absorbencies were analysed at 360 nm using an automated microplate reader (Elx800-Universal Microplate Reader, BioTek® Instruments, Inc.).

Energy consumption Cellular respiration and the rate thereof, was measured within Synodontis zambezensis through the analysis of ETS activity (De Coen and Janssen, 1997, 2003). Homogenised samples were centrifuged (High Speed Refrigerated Centrifuge: NovaFuge B113-21R, Senova Biotech Co.,Ltd) at 4˚C, 2 940 g. for 10 mins. The procedural blank used for this protocol was ETS buffer (pH 8.5) at a volume of 25 µL. The same volume of each sample was used and to the blanks and the samples 75 µL Buffered substrate solution (BSS) (as per protocol), 25 µL NADPH and 50 µL 3-(4-Iodophenyl)-2-(4-nitrophenyl)-5-phenyl-2H-tetrazol-3-ium chloride (INT, 95%, Sigma-Aldrich) was added. The analysis of the plate was done kinetically at 490 nm, 20˚C for a total of 5 mins at 1 min intervals using an automated microplate reader (Elx800-Universal Microplate Reader, BioTek® Instruments, Inc.).

The enthalpy of combustion values (De Coen and Janssen, 1997) were used to convert energy reserves into energy equivalents. The equation used for the calculation of total CEA is as follows:

퐶퐸퐴 = 퐸푎 − 퐸푐

Where 퐸푎 = 퐸푔푙푢푐표푠푒 + 퐸푙푖푝푖푑 + 퐸푝푟표푡푒푖푛 and 퐸푐 = 퐸퐸푇푆

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6.1.3 Statistical analysis SPSS version 25 (IBM Software Group) was used for one-way analysis of variance (ANOVA). The Chi-square test was performed on all data to ensure homogeneity following which, all biomarker and bio-accumulation data was log transformed.. Significant differences (p≤0.05) between concentrations and skewness of data were determined using the Tukey’s Post Hoc test (Yohannes et al., 2013). Significant differences (p≤0.05) in the same biomarker, between tissues (AChE in muscle vs. AChE in liver) were determined in Excel using a student’s t-test since all data was found to be normally distributed (Hayes and Kruger, 2014). Multivariate statistics were applied to the bio-accumulation and biomarker data. Constrained redundancy analysis (RDA) was applied to determine the influence of 4,4’-DDT and its degradation products on the spread of the exposure biomarkers analysed (Quinn and Keough, 2002). The constrained RDA was completed with the use of Canoco 5 software.

6.2 Results 6.2.1 Biomarkers of exposure and energetics vs. 4,4’-DDT The AChE absorbance concentrations in the muscle tissue range from a mean concentration of 0.004 abs/min/mg protein to 0.01 abs/min/mg protein, and in the liver tissue from a mean of 0.06 abs/min/mg protein to 0.07 abs/min/mg protein (Table 6.1). The results for AChE in the liver tissue can be seen in both Figure 6.1A and 6.1B. In Figure 6.1A we see a significantly higher concentration (p≤0.05) of AChE at the 31.58 ng/mL exposure concentration when compared to the 15.79 ng/mL exposure concentration. This is an unexpected result as AChE concentrations should be inhibited, therefore lower AChE concentrations as 4,4’-DDT exposure concentrations increase. Observations made in Figure 6.1B indicate significant differences (p≤0.05) in AChE between the muscle and liver tissue at the two specific exposure concentrations. At both the 3.95 ng/mL and 31.58 ng/mL exposure concentrations, there are significantly lower (p≤0.05) concentrations of AChE found in the muscle tissue when compared with the concentrations found in the liver tissue. This shows the inhibitory effects 4,4’-DDT is having on the AChE within the muscle tissue when compared to the excitatory effect it seems to be having in the liver tissue, although this was not statistically significant. The overall trend seen in Figure 6.1B is the general inhibition of AChE in the muscle tissue as the exposure concentrations increase.

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In Figure 6.1A there is only one observable significant difference (p≤0.05) noted in the CYP450 concentrations, with the control having significantly higher concentrations than those found at the 3.95 ng/mL exposure concentration. This is an unexpected result, as an excitatory response is a general expectation as the exposure concentrations increase. The CYP450 concentrations range from a mean concentration of 0.0004 nM/mg protein to 0.0005 nM/mg protein (Table 6.1).

Table 6. 1: Resulting concentrations of AChE and CYP450 (Mean ± SD) and range in the liver tissue of Synodontis zambezensis after an acute exposure to 4,4’-DDT. Significant differences (p≤0.05) per column are indicated by common superscripts in the form of capital or lower case letters. AChE Conc. C-P450 (abs/min/mg (ng/mL) (nM/mg protein) protein) a0.006 ± 0.0005 0.0004 ± 0.00002 Control (0.004 - 0.007 (0.0004 - 0.0005) 0.005 ± 0.0004 0.0005 ± 0.00002 1.97 (0.004 - 0.007) (0.0004 - 0.0006) a0.005 ± 0.0002 0.0005 ± 0.00001

3.95 (0.005 - 0.006) (0.0004 - 0.0005)

Liver 0.006 ± 0.0002 0.0005 ± 0.00001 7.90 (0.005 - 0.006) (0.0004 - 0.0005) 0.007 ± 0.0007 A0.0005± 0.00001 15.79 (0.004 - 0.01) (0.0004 - 0.0005) 0.004 ± 0.0003 A0.0005 ± 0.00003 31.58 (0.003 - 0.005) (0.0004 - 0.0006) 0.005 ± 0.00009 0.0004 ± 0.00001 63.17 (0.004 - 0.005) (0.0003 - 0.0004)

Figure 6.1C and 6.1D illustrate the resulting biomarker reactions for overall energetics (CEA) and energy availability (Ea) respectively. Beginning with total CEA a significantly higher (p≤0.05) concentration is observed in both the control and the 63.17 ng/mL exposure concentrations when compared to the 15.79 ng/mL exposure concentration. This indicates that CEA drops in the lower exposure concentrations, but then increases again as the exposure concentrations increase. Total CEA ranges from a mean concentration of 350.73 J/g to 544.18 J/g (Table 6.2). The Ec concentrations range from a mean concentration of 160.81 J/g to 309.45 J/g (Table 6.2) with significantly higher (p≤0.05) concentrations at all of the exposure concentrations when compared to the control, showing an increase in the energy consumed by the fish as the exposure concentrations increase. The available energy (Ea) within Synodontis zambezensis (Figure 6.1D) remains fairly constant from the

81 | P a g e control throughout the increasing exposure concentrations and ranges from a mean concentration of 627.45 J/g to 780.21 J/g (Table 6.2).

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Table 6. 2: Resulting concentrations of AChE, CEA, Ec, Ea, glucose, lipids and proteins (Mean ± SD) and range in the muscle tissue of Synodontis zambezensis after an acute exposure to 4,4’-DDT. Significant differences (p≤0.05) per column are indicated by common superscripts in the form of capital or lower case letters. AChE Conc. CEA Ec Ea Glucose Lipids Protein (abs/min/mg (ng/mL) (J/g) (J/g) (J/g) (J/g) (J/g) (J/g) protein) 0.004 ± 0.0005 A507.72 ± 17.53 abcdef160.81 ± 11.88 668.53 ± 11.15 ABC9.43 ± 0.74 227.71 ± 7.91 431.39 ± 13.02 Control (0.0003 - 0.005) (418.78 - 602.59) (72.24 - 216.73) (635.52 - 761.23) (4.95 - 14.41) (187.83 - 262.51) (390.80 - 523.20) 0.007 ± 0.01 413.41 ± 27.55 a240.30 ± 12.37 653.70 ± 23.40 D13.35 ± 2.68 224.60 ± 19.79 415.75 ± 8.58 1.97 (0.003 - 0.01) (311.37 - 605.25) (178.23 - 300.46) (556.17 - 829.41) (3.15 - 29.27) (155.67 - 389.47) (360.00 - 449.47)

0.009 ± 0.0008 376.73 ± 32.13 b294.86 ± 6.39 671.59 ± 29.68 AD32.99 ± 7.50 227.71 ± 17.15 410.89 ± 12.30 3.95 (0.006 - 0.01) (273.87 - 638.75) (265.44 - 325.62) (599.50 - 921.01) (14.86 - 92.32) (160.56 - 353.36) (336.27 - 475.33)

Muscle 0.01 ± 0.0004 435.55 ± 48.20 c288.20 ± 7.77 723.76 ± 44.43 B32.67 ± 8.88 275.07± 32.65 416.01 ± 9.42 7.90 (0.008 - 0.01) (225.61 - 730.97) (252.92 - 338.15) (537.90 - 942.55) (7.43 - 93.44) (187.27 - 507.22) (343.20 - 450.00) 0.01 ± 0.0008 AB350.73 ± 31.66 d309.45 ± 11.68 660.17 ± 30.85 20.04 ± 2.66 241.44 ± 26.82 398.69 ± 7.45 15.79 (0.006 - 0.01) (205.06 - 535.16) (250.13 - 376.30) (580.52 - 816.13) (11.26 - 32.65) (174.67 - 399.63) (365.20 - 447.47) 0.01 ± 0.001 393.57 ± 33.23 e233.89 ± 18. 45 627.45 ± 26.90 C29.74 ± 6.32 203.39 ± 15.08 a394.32 ± 9.81 31.58 (0.006 - 0.02) (243.02 - 546.88) (159.91 - 350.67) (538.19 - 814.98) (7.88 - 76.78) (147.02 - 292.60) (328.93 - 445.60) 0.01 ± 0.001 B544.18 ± 80.81 f236.03 ± 19.85 780.21 ± 84.50 16.84 ± 3.58 340.75 ± 73.44 a422.61 ± 23.67 63.17 (0.010 - 0.013) (310.57 - 780.19) (189.37 - 291.99) (574.50 - 1072.18) (4.05 - 24.99) (197.24 - 607.10) (352.27 - 482.40)

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However, when looking at the individual components of Ea a significantly higher (p≤0.05) glucose concentration is seen at the 3.95 ng/mL, 7.90 ng/mL and 31.58 ng/mL exposure concentrations when compared to the control. There is also a significantly higher (p≤0.05) glucose concentration found at the 3.95 ng/mL exposure concentrations when compared with the 1.97 ng/mL exposure concentration. There is a general increasing trend in glucose concentrations as the exposure concentrations increase, and the glucose concentrations range from a mean of 9.43 J/g to 32.99 J/g (Table 6.2).

Figure 6. 1: Biomarkers of exposure (mean and standard deviation (SD)) within the liver (A), AChE in the liver and muscle tissue (B), energetics (C) and Ea and its components (D) in Synodontis zambezensis after an acute exposure to various concentrations of 4,4’-DDT. Significant differences (p≤0.05) are indicated by the presence of common superscripts (capital letters or lower case letters). Resulting biomarker concentrations have been log transformed in order for them to be illustrated together on the corresponding graphs.

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There are no significant differences (p≤0.05) seen in the protein concentrations throughout the exposure concentrations, but there is a significantly higher lipid concentrations found in the 63.17 ng/mL exposure concentration when compared to the 31.58 ng/mL exposure concentration. The protein concentrations range from a mean of 394.32 J/g to 431.39 J/g and the lipid concentrations range from a mean of 203.39 J/g to 340.75 J/g (Table 6.2). There is again an increasing trend in concentrations of lipids as the exposure concentrations increase. The increasing lipids and glucose within S. zambezensis does not affect the overall Ea, but shows an exposure response to 4,4’-DDT.

In order to relate the biomarkers of exposure and energetics to 4,4’-DDT and its degradation products, an RDA triplot was constructed (Figure 6.2). The biomarkers of exposure and energetics analyses were constrained by the concentrations of 4,4’- DDT and its degradation products. This constrained ordination gives an indication of the effect of 4,4’-DDT and its degradation products on the spread of the biomarkers of exposure and energetics.

The ordination in Figure 6.2 explains 96.18% of the variance within the data. The first axis explains 70.47 % of the variance within the data. The second axis explains 25.71 % of the variance within the data and a clear separation is seen vertically between AChE in the muscle tissue, CEA, Ea lipids when compared to AChE in the liver, Ec, CYP450, protein and glucose. The arrow for 4,4’-DDT is much longer than that of its degradation products and this indicates its increased influence on the resulting biomarker excitation.

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Figure 6. 2: An RDA triplot showing resulting biomarkers of exposure and energetics in Synodontis zambezensis following an acute exposure to 4,4’-DDT.

The angle between CEA, lipids and Ec (<90˚ angle) indicates a strong correlation of these biomarkers with 4,4’-DDD and 4,4-DDE. Glucose, protein and CYP450 all show a weak correlation (≥90˚ angle) with 4,4’-DDT and its degradation products. There is however a strong correlation between 4,4’-DDT and its degradation products with AChE in the liver and a week correlation with AChE in the muscle. This is because this biomarker acts inversely to the other biomarkers since it is inhibited and not excited by the action of toxins (Sturm et al., 1999), what would normally be seen as a strong correlation as in AChE in the muscle (<90˚ angle) is in fact a weak correlation and vice versa.

6.3 Discussion 6.3.1 Biomarkers of exposure and energetics vs. 4,4’-DDT Biomarkers of exposure are generally used in field studies to indicate possible exposure of organisms to chemicals in their surrounding environments (Van der Oost et al., 2003). In this case Synodontis zambezensis were exposed to 4,4’-DDT in a controlled laboratory environment and the resulting biomarker responses are expected to reflect the increasing concentrations of 4,4’-DDT. In the liver biomarkers,

86 | P a g e when referring to Figure 6.1A, an almost constant CYP450 concentration is observed from the control to the highest concentration, with only one significant difference (p≤0.05) present. This is an unexpected result, since the CYP450 superfamily is responsible for phase 1 detoxification of pollutants (Andersson and Förlin, 1992). This unexpected result could have occurred because the exposure period was acute and not chronic. The time required for many biomarker processes to show notable physiological changes may exceed 96 hours and for this reason a false negative may have resulted in this study (Forbes et al., 2006). The same can be said for the relative stability of AChE in the liver tissue (Figure 6.1A) throughout the increasing exposure concentrations. The comparison between AChE in the liver and muscle tissue shows an interesting trend with the AChE in the muscle tissue appearing to decrease as the exposure concentrations increase. This is an expected result as the inhibition of this biomarker is an indication of increased exposure to pollutants (Leibel, 1988). When observing the trends in the RDA, however, it is noted that AChE in the liver is positively correlated with the effects of 4,4’-DDT and not AChE in the muscle tissue. This may be the case since for the RDA ordination the effects of each individual within an exposure concentration have been taken into consideration and not just an average result as seen in the graphs (Figure 6.1B).

The total CEA is clearly affected by the increasing concentrations of 4,4’-DDT, but again not in the manner that was expected. Metabolism and energy utilization can be greatly affected by the increasing presence of pollutants (De Coen and Janssen, 1997), but it is more useful in this case to look at the different components of CEA. The Ea levels appear to remain relatively stable throughout the exposure period, but they are higher than the Ec at every exposure concentration, indicating a trend towards the storage of energy, rather than increased metabolism. When focussing on the constituents of Ea (glucose, lipids and proteins) apparent changes are visible. Glucose concentrations steadily increase as the exposure concentrations increase and then begin to drop again, in a sort of bimodal reaction. There are no significant differences (p≤0.05) observed in the protein concentrations, and only one observed in the lipid concentrations. The fluctuations in glucose may be related to the cortisol- glucose pathway within teleost fishes (Barton, 2002).

Generally cortisol regulates the storage and release of glucose for use in the body, and under stressful conditions increased blood glucose concentrations would

87 | P a g e indicate activation through stress (Barton, 2002). Since the muscle tissue and not the blood plasma was tested in this instance, it is unclear whether there was an increase in blood glucose as the tissue glucose concentrations decreased again in the higher exposure concentrations.

6.4 Conclusions and recommendations It is clear from these results that a response is taking place within S. zambezensis but because the exposure was acute and not chronic it is unclear whether the responses being recorded are true indications of physiological adaptation (Forbes et al., 2006). These results are every important for future exposure studies using S. zambezensis as the test organism as they may be used as baseline responses to 4,4’-DDT under acute conditions. From these results chronic exposure studies are recommended as well as a rerun of the acute exposure experiments with the addition of other biomarkers which were not the focus in this study. Biomarkers such as heat shock proteins (HSP), blood cortisol and lactate levels should be considered for future analyses. A recommendation would also be to introduce time-specific biomarker analysis protocol. This would mean the subsampling of S. zambezensis from each concentration at 24 hour intervals. Both accumulation and biomarkers could then be analysed every 24 hours, giving a clearer indication of the accumulation and the physiological responses linked directly to time.

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Chapter 7: Ribonucleic acid (RNA) as a biomarker

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7. Introduction In gene expression studies, RNA is often extracted and subsequently converted to complimentary DNA (cDNA) (Trieu et al., 2014). The changes in gene expression due to exposure to external toxicants can be mapped through the use of the techniques outlined in this chapter. After extraction of RNA as well as post conversion to cDNA, the initial screening technique used to assess the presence of target genes is through gel electrophoresis following polymerase chain reaction (PCR) with suitably designed primers.

Polymerase chain reaction (Figure 7.1) has become an important and powerful tool in molecular biology, beginning with a single molecule of DNA and ending in hundreds of similar molecules (Mullis, 1990). The ease of execution and limited reagents and equipment required to complete PCR makes it an attractive tool in many laboratories (Mullis, 1990). The amplification and exponential duplication of specific fragments of DNA has limitless applications and has become indispensable in the medical and biological fields (Solanki, 2012).

Following a basic principle, PCR is a chain reaction of multiple heating and cooling cycles (Solanki, 2012), beginning with a single DNA molecule. The exponential duplication of a single DNA molecule to hundreds or thousands more, is accomplished by proteins known as polymerases (Chien et al., 1976). Polymerases initiate the formation of a new polypeptide chain by adding the complimentary nucleotide bases to the template strand of DNA (Berg et al., 2002). In PCR a specific polymerase is used since it is required to withstand temperatures as high as 94˚C without denaturation (Chien et al., 1976). This DNA polymerase was extracted from extreme thermophilic eubacteria, Thermus aquaticus, and is known as Ex Taq in PCR amplification kits (Chien et al., 1976). The polymerase “knows” where to start adding the new nucleotide base pairs based on a supplied primer (Chen and Janes, 2002). These primers may be specific to a section of the DNA and the detail on their design will be discussed later in this chapter (Dieffenbach et al., 1993). The building blocks required for the duplication of DNA are the four nucleotides namely adenine (A), thymine (T), guanine (G) and cytosine (C), which are supplied in the reaction by deoxy-ribonucleotide triphosphates (dNTPs) (Lodish et al., 2000). The addition of the new nucleotide bases will be completed through the action of the polymerase and will follow the standard A=T, G≡C bonding in the growing DNA strand (Lodish et al.,

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2000). The final requirement for a successful PCR reaction is a buffer that allows the reaction to take place within a stable pH range and curbs any possible denaturation due to acidity (Solanki, 2012). The buffer used is known as 10xEx Taq Buffer solution. It contains Mg2+ and is essential for the optimal activity of the Ex Taq polymerase (Chien et al., 1976).

Primers are defined as short nucleic acid fragments that are single-stranded (oligomers) (Chen and Janes, 2002) and correspond with a desired target DNA (Chen and Janes, 2002). They are used in PCR for amplification and logarithmic duplication of desired sections of DNA (Singh and Kumar, 2001). Each set of primers consist of a forward and a reverse primer, allowing for the amplification and duplication of a segment of DNA which contains a target gene (Chen and Janes, 2002). The forward primer direction is the same as that of the target DNA strand and runs from 5’ to 3’, but the reverse primer is designed initially in the 3’ to 5’ direction and then changed into a reverse compliment in order to run in the 5’ to 3’ direction (Singh and Kumar, 2001). Primers are designed to be specific and efficient and in order for them to meet these two criteria they need to be designed within a few fundamental guidelines (Dieffenbach et al., 1993). The size of each primer (forward and reverse) should ideally be between 15 and 25 nucleotides (Chen and Janes, 2002). If the primer is too short it may not be specific enough to magnify the target section and a primer that is too long may have a reduced ability to anneal to the target DNA (Dieffenbach et al., 1993). The G and C content of primers should be between 50 and 60% to assure stability of the primer (Chen and Janes, 2002), offered by the triple bonding found between these two amino acids (Yakovchuk et al., 2006). Stability is further offered by ensuring that G and C are present in at least the last three bases at the 5’ terminal, creating a GC clamp (Sheffield et al., 1989). A GC clamp should be avoided at the 3’ end of the primer as this may cause miss- priming during PCR (Huang et al., 1992). The melting temperature (Tm) of a primer is the temperature at which 50% of the designed forward or reverse primer molecules are bound to the template DNA and the other 50% exists dissolved in solution (Dieffenbach et al., 1993). The melting temperature of primers is directly proportional to the GC content (Marmur and Doty, 1962), the higher the GC content, the higher the Tm. This is an important consideration when designing applicable primers because the forward and reverse primer Tm’s should be within 5˚C of one

91 | P a g e another to allow for simultaneous use in a PCR reaction (Dieffenbach et al., 1993). The optimal range for Tm of the forward and reverse primers is between 56˚C and 62˚C (Dieffenbach et al., 1993).

Secondary structures, which may occur when designing primers include hairpin loops and primer dimers (Singh et al., 2000). These secondary structures can lead to a decrease in the efficiency of the primers to amplify the target sequence (Singh et al., 2000). Hairpin loops occur when the primer is self-complimentary, containing runs of nucleotides which complement one another as well as the target sequence (Innis and Gefland, 1990). This self-complimentary section may result in the folding back of the primer (Chen and Janes, 2002). The inefficiency caused by hairpin loops is a result of the unavailability of sections of the primer for amplification of the target sequence (Chen and Janes, 2002; Singh et al., 2000). Primer-dimer formation occurs when the forward and reverse primers are complimentary to one another and the effect on amplification is the same as that seen during the formation of hairpin- loops (Singh et al., 2000).

The three steps in every PCR reaction are denaturation, annealing and extension (Solanki, 2012). The first phase begins with the heating of the PCR components to temperatures between 90˚C and 97˚C, causing the DNA double-helix to break apart into two separate strands (Wang et al., 2014). Now that the sections of DNA are available for duplication and amplification the temperature is lowered in the reaction tube and the primers are able to anneal or bond with the target section of the single DNA template strand (Solanki, 2012). The temperature of this step is determined by the Tm of the primer pair (Singh and Kumar, 2001). The final step is the elongation of the complimentary DNA strand starting at the 3’ end of the forward and reverse primers moving in the 5’ to 3’ direction (Singh and Kumar, 2001). The temperature of this step is determined by the optimum activity temperature of Thermus aquaticus (Ex Taq) which is between 70˚C and 80˚C (Chien et al., 1976). Final elongation is achieved at a temperature of 72˚C for 4 to 15 mins before final decrease of the temperature to 4˚C for storage of PCR product (Chen and Janes, 2002).

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Figure 7. 1: Polymerase Chain Reaction steps and components (Chou et al., 1992).

The formation of recombinant DNA through the process known as molecular cloning, is commonly used, alongside PCR to replicate specific DNA sequences (Watson, 2007). While PCR takes place in a PCR tube without the presence of any living cells, molecular cloning takes place within a living cell (Watson, 2007). A cloning vector is required for the formation of recombinant DNA (Sambrook et al., 1989). Vectors are bacterial DNA molecules known as plasmids which will carry the gene of interest (Sambrook et al., 1989). A cloning vector is used during this process that allows for the insertion of foreign DNA and subsequent transformation of this vector into a suitable bacterial cell for replication (Watson, 2007). Cloning is most often performed using Escherichia coli (E. coli) as the competent cell because of its ability to take up “naked” DNA (Sambrook et al., 1989). Once the plasmids have been transformed into the E. coli bacteria, replication of the desired sequence takes place and the bacterial cells are able to produce millions of copies of the desired sequence within a couple of hours (Watson, 2007). The replication of bacterial colonies takes place in Lysogeny Broth (LB), which is prepared as a medium for growth in the form of a simple solution or used to prepare the Petri dishes used during blue/white screening techniques (Bertani, 1951).

Lysogeny Broth was first developed by Guiseppe Bertani for the growth of Shigella sp. bacteria (Bertani, 1951). This medium is rich in nutrients required for bacterial growth and has been used predominantly for culturing E.coli and other strains of gram-negative bacteria (Anderson, 1946; Bertani, 2004; Luria et al., 1960; Luria and

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Burrous, 1957). The ingredients that make up this culture medium include tryptone, yeast extract, sodium chloride and sodium hydroxide (Bertani, 1951). Tryptone is a source of amino acids in LB medium and is formed through the digestion of casein by trypsin (Fraser and Powell, 1950). Yeasts and yeast extracts have been used for many years in fermentation processes of beer, wine and bread (Bekatorou et al., 2006). Baker’s yeast, as the name suggests, is used commercially in the baking of breads (Bekatorou et al., 2006). Brewer’s yeast is used commercially in the brewing industry (Bekatorou et al., 2006) and yeast extracts are used for various forms of bacterial growth medium such as LB medium (Chae et al., 2001). The manufacture of yeast extract is achieved through the breakdown of yeast cells using various enzymes (Choi and Chung, 1998). Yeast extract provides amino acids, peptides and nucleotides to bacteria cultured in LB medium (Chae et al., 2001). The addition of NaCl and NaOH to the medium provides a low salt content and aids in the growth of bacterial colonies (Bertani, 1951).

Once transformation and cloning has taken place, the successful bacterial colonies need to be identified with the naked eye; this is done through blue/white screening (Padmanabhan et al., 2011). Once transformed, the Douglas Hanahan 5 alpha (DH5α) competent cells (E. coli) are allowed to grow in an agar Petri dish in the presence of 5-bromo-4-chloro-1H-indol-3-yl β-D-galactopyranoside (X-gal) (Padmanabhan et al., 2011). Successful colonies will be white in colour and will contain recombinant DNA. Unsuccessful colonies will be light to dark blue in colour (Ullmann et al., 1978) (Figure 7.2).

Figure 7. 2: Figure showing the presence of both white (successful) and blue (unsuccessful) bacterial colonies (Padmanabhan et al., 2011).

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Gel electrophoresis is an effective way to separate out fragments of DNA or in this case cDNA following PCR amplification or molecular cloning, and can be applied to fragments ranging in size from 100 bp to 25 kbp (Sambrook et al., 1989). The ingredients required to make up an electrophoresis gel include agarose and tris- borate-ethylnenediaminetetraacetic acid (EDTA) buffer (TBE Buffer) (Lee et al., 2012). Agarose is a powder extracted from seaweed in the genera Gelidium and Gracilaria and is added to the TBE buffer based on the size of the DNA fragments you wish to separate (Kirkpatrick, 1990). Agarose gels are described in terms of percentages and therefore the percentage gel you would like to use and make is calculated before each gel solution is made. The percentage gel required for separation depends on 1) the size of the fragments separated, 2) the specificity of the separation and 3) whether or not “gene cleaning” of the gel slices is required after separation. The larger fragments of DNA and/or RNA are, the lower the percentage of agarose, since the pore sizes of the gel decrease proportionally as the percentage of agarose increases (Lee et al., 2012). When calculating the gel percentage the calculation applied is a mass to volume ratio.

In order for successful comparisons on gene expression to be made, it is essential to first identify a “housekeeping” (HK) gene within the target species. These genes are usually constitutive genes that are essential for the normal functioning of cells (Eisenberg and Levanon, 2003). These genes are present and are expressed under normal conditions, as well as under pathophysiological conditions and for this reason they are used as baselines for the expression of other genes within the organism (Butte et al., 2001; McCurdy et al., 2008). The most common housekeeping genes used in comparative expression studies are beta-actin (β-actin), glucose-6- phosphate dehydrogenase (g6pd) and glyceraldehyde 3-phosphate dehydrogenase (gapdh) (McCurley and Callard, 2008). The HK gene chosen for this study was β- actin because it has been used in previous studies with fish from the family Siluriformes (García-Reyero et al., 2004; Huang et al., 2011; Kim et al., 2012; Wu et al., 2009). Actins are proteins present within animal cells which are highly conserved and play major roles in cell motility and structural integrity (Gunning et al., 2015). In a study by McCurley and Callard (2008), β-actin was shown to be highly expressed in the tissues (eye, brain, heart, liver, muscle and gonad) of the Zebrafish (Danio rerio) when compared to other common HK genes.

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The other genes investigated for expression analysis were CYP1A, transferrin, vitellogenin I and II (VTG I and VTG II), oestrogen receptor (ER) and choriogenin H and L (CHg H and CHg L).

The effects of environmental contamination are often shown through the analysis of Cytochrome P450 (Hahn et al., 1998). The proteins in the CYP group play an important role in the oxidative metabolism of compounds found within animals as well as toxic compounds introduced from the external environment, which may include pharmaceuticals (Diotel et al., 2010). Vertebrates all have the known CYP1 subfamilies (CYP1A, CYP1B, CYP1C and CYP1D) with mammalian vertebrates containing only CYP1A and CYP1B (Goldstone et al., 2007). While mammals have two CYP1A genes (CYP1A1 and CYP1A2), fish species only have one CYP1A gene (Goldstone et al., 2007).

Transferrin is one of the four plasma proteins, present in all vertebrates, responsible for the transport of iron around the body and the prevention of loss of iron from the blood through urinary excretion (Putnam, 1975). Although its role in the blood is linked to other blood proteins, transferrin is specifically involved in the transport of iron to the “red” bone marrow (erythropoietic bone marrow) (Weinberg, 1978). Iron transport is also mediated to the reticuloendothelial system (spleen, liver, small intestine and muscle) through the action of transferrin, making this protein instrumental in the immune system and production of red blood cells of vertebrate species (Weinberg, 1978).

Since 4,4’-DDT is a known endocrine disrupting chemical (EDC) (Leaños-Castañeda et al., 2007), the Chg (H and L), VTG (I and II) and ER genes were chosen to show any effect the exposure may be having on the endocrine system of the target species S. zambezensis. Choriogenins (Chgs) are synthesized in the liver of fishes and transported to the ovaries via the blood stream (Arukwe and Goksøyr, 2003; Murata et al., 1995; Yamagami, 1996). They are produced in response to oestrogens and are precursors of the inner layer of the egg envelope (Hamazaki et al., 1989; Litscher and Wassarman, 2007). The “H” in choriogenin H represents the “higher” relative molecular mass of this glycoprotein compared to the “lower” molecular mass of Chg L (Murata et al., 1995). Fish Chg genes are regarded as sensitive biomarkers of exposure to EDCs (Lee et al., 2002) and are regulated through oestrogen

96 | P a g e responsive elements (ERE) (Ueno et al., 2004). Vitellogenins are synthesized in the liver and transported via the bloodstream to the ovary (Murata et al., 1995) where they are absorbed into the growing oocytes, this is known as vitellogenesis and forms an important part of oogenesis (Arukwe and Goksøyr, 2003). These VTGs are protein precursors to the yolk found in developing eggs of fish, birds and amphibians (Wallace, 1985) and are also produced under the stimulation of oestrogen (Hamazaki et al., 1989). Since one of the common effects of EDCs is the mimicking of oestrogens, VTGs can also be used as important biomarkers of exposure to EDCs (Arukwe and Goksøyr, 2003). The expression of both Chgs and VTGs are influenced and mediated by oestrogen binding to oestrogen receptors (ER) (Beato, 1989). Vitellogenin and Chgs are normally produced in the liver of mature female fish, but male fish that are affected by internal or external oestrogen or EDCs can also produce this protein precursor in their livers (Mommsen and Walsh, 1988). Chemicals that bind to ER have the ability to activate transcription of oestrogen dependent genes like VTGs and Chgs (Flouriot et al., 1995; Petit et al., 1999), resulting in excess levels of these genes.

The importance of the quantification of nucleic acids for diagnostics and biomedical research has led to the development of various technologies (Lekanne Deprez et al., 2002). Previous PCR-assays (Gibson et al., 1996) have been vastly improved through the use of fluorescent detection (Rasmussen, 2001). Quantitative real-time PCR (qRT-PCR) allows for real-time detection of PCR products and also gives an indication of the efficiency of the primer used, since the data is collected in a log linear phase (Rasmussen, 2001). The amount of cellular RNA gives an indication of the expression of a particular gene (Kim et al., 2008). SYBR® Green is a fluorescent dye that binds to double stranded DNA (Ponchel, 2006). As the PCR protocol continues the amount of double stranded DNA also increases and this is reflected by the increase in fluorescence expressed through SYBR® Green (Ponchel, 2006). There are two methods for quantifying the levels of the expressed genes namely absolute or relative quantification (Pfaffl, 2006). In absolute quantification, the PCR signal is related to input copy number and is based on a calibration curve (Bustin, 2000). This method can be misleading since it is based on a calibration curve and allows for increased human error when setting up the calibration curve (Pfaffl, 2001). Relative quantification determines the changes in gene expression across multiple

97 | P a g e samples in relation to levels of another gene and does not require the use of a calibration curve (Bustin, 2002). The genes chosen for comparison are the HK genes, which show constant expression across a variety of genes and under various conditions (Dheda et al., 2004). Before comparisons of gene expression can be completed, it is important to establish the efficiency of the primers that have been sequenced to amplify and replicate the genes of interest (GOI) (Pfaffl, 2006). This can be done through the dilution method and should be completed for each sample forming part of an experiment (Souazé et al., 1996). This method can therefore be very time consuming and laborious (Souazé et al., 1996).

The aim of this section of the study was to isolate RNA from liver, kidney, muscle and gonadal tissue of S. zambezensis. Following RNA isolation conversion to cDNA was completed and the designing and sequencing of successful primers was completed. Primer efficiencies were calculated based on resulting qRT-PCR data. From this point the aim was to establish whether the expression of the target genes had been affected by the exposure to varying concentrations of 4,4’-DDT. The objectives used to complete these aims will be outlined during the materials and methods sections to follow.

7.1 Materials and methods 7.1.1 Fish tissue collection After the 96-hour exposure period, fish were killed by severing the spinal cord and the liver, kidneys and gonads were removed and placed into RNAlater (Sigma- Aldrich) to reduce RNA degradation. The samples were placed into a -80˚C freezer and stored until RNA extraction and analysis was completed.

7.1.2 RNA isolation reference The tissue samples were removed from the freezer and allowed to thaw. A small section (±50 mg) of each tissue sample was placed into an Eppendorf tube containing a ceramic homogenizing bead (United Scientific (Pty) Ltd.) and 700 µL TRIzol Reagent (Sigma-Aldrich) (Hummon et al., 2007). The tissue was homogenized using a Qiagen tissueLyser bead homogenizer (Qiagen, South Africa). Once completely homogenized, the homogenate was pipetted into a new Eppendorf tube and 200 µL of chloroform was added. The tubes were vortexed to ensure homogeneity of the mixture and centrifuged at 4˚C, 1200 g for 20 min. After

98 | P a g e centrifugation a clear separation is visible within the tubes with three visible layers (Figure 7.3), the top layer is the required RNA layer (Hummon et al., 2007).

Figure 7. 3: Resulting separation after the centrifugation during TRIzol RNA extraction (Hummon et al., 2007).

From this point on the Nucleospin®RNA isolation kit was used (Macherey-Nagel, 2015), and the manufacturer’s instructions were followed. The wash buffer RA3 and rDNase solutions were prepared before using this isolation kit. The protocol began at step 4 by pipetting the supernatant directly onto the filter of a new blue-ring filter tube (Figure 7.4).

Figure 7. 4: The blue-ring filter tube supplied in the Nucleospin®RNA isolation kit (Macherey- Nagel, 2015).

This was followed by pipetting 350 μL 70% ethanol directly onto the filter and centrifuging at 4˚C, 11 000 g for 30 sec. The samples were removed from the centrifuge and 350 μL membrane desalting buffer (MDB) was added directly to the filter and this was again centrifuged at 4˚C, 11000 g for 1 min. Ninety five μL of rDNase was added to each filter and the tubes were left to incubate at room temperature (±25˚C) for 15 mins. After the incubation period, the tubes were placed on ice and 200 μL RA2 Buffer was added and centrifuged at 4˚C, 11000 g. for 1 min.

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The resulting solution was discarded, the filter was removed and placed onto a new collection tube and 600 μL RA3 Buffer was added and centrifuged at 4 ˚C, 11000 g. for 1 min. This was followed by the addition of another 250 μL RA3 buffer and centrifugation at 4 ˚C, 11 000 g. for 2 min. The resulting solution was discarded and the filter was transferred again, this time to a 1.5 mL nuclease-free collection tube. The RNA was eluted from the filter with 60 μL RNase free water and centrifugation at 4˚C, 11000 g. for 1 min followed.

Once isolation was completed, the resulting RNA quality and quantity was checked using a Nanodrop® ND-1000 spectrophotometer (Thermo Fisher Scientific). The ratio of the absorbance at 260 nm and 280 nm was used to assess the purity of the RNA. The 260/280 readings as well as the concentration in ng/μL were recorded for each sample. The 260/280 readings are required to be between 1.8 and 2 for good quality samples. All of the isolated RNA fell within this range.

After quality control was completed, the isolated RNA was run on a 1% agarose gel to confirm the presence of RNA. The presence of RNA was confirmed by the presence of 18s and 28s bands on the completed gel (Imbeaud et al., 2005).

7.1.3 cDNA synthesis After RNA isolation, cDNA synthesis was completed. In order to perform PCR on the RNA samples to acquire cDNA, the protein concentration of each RNA template needs to be between 0.5 pg and 0.5 μg. For this reason, all of the RNA samples were diluted based on the protein concentrations obtained from the Nanodrop® ND- 1000 Spectrophotometer (Thermo Fisher Scientific). cDNA synthesis was completed with the use of the ReverTra Ace® qPCR RT master mix with genomic DNA (gDNA) remover 1204 (Toyobo, Japan) synthesis kit. Before beginning the procedure, the 4xDN master mix and gDNA remover mixture was made and stored at -20˚C. The whole procedure was completed on ice. The first stage of the cDNA synthesis was the denaturing of the RNA template. This was completed by pipetting a relevant amount of the RNA template (this depends on protein concentration as explained above) into a 0.2 mL polypropylene PCR Tube (Strip containing 8 tubes). The tube strip was placed into the Bio-Rad iCycler PCR System (Bio-Rad Laboratories) at 65˚C for 5 min. The strip was removed, placed on ice and 10 μL 4xDN master mix and 4 μL nuclease-free water was added. This

100 | P a g e mixture was incubated at 37˚C for 5 min for gDNA Removal. After the incubation period, the strip was again placed on ice and 10 μL 5xRT master mix II was added. This final mixture was placed into the Bio-Rad iCycler PCR System (Bio-Rad Laboratories) for final cDNA synthesis (reverse transcription). The cycle was as follows: 37˚C for 15 mins, 50˚C for 5 mins and heated to 98˚C for 5 min.

The final quality and protein quantity of the resulting cDNA was checked again using a Nanodrop® ND-1000 Spectrophotometer (Thermo Fisher Scientific).

7.1.4 Primer design protocol The initial stage in primer design was to identify the desired genes for amplification. Nucleotide sequences were accessed using the National Centre for Biotechnology Information (NCBI) database. Genes specific to Synodontis zambezensis were searched and since there was a lack of data during the completion of this section of the study, gene sequences from the family Siluriformes were also included as part of the search. The FASTA format (Lipman and Pearson, 1985; Pearson and Lipman, 1988) of the gene sequences where extracted from NCBI and alignment of the sequences was completed using gene alignment software, Clustal W (version 2.0).

Figure 7. 5: Gene alignment example (Larkin et al., 2007), with an asterisk (*) indicating a single fully conserved residue, a colon (:) indicating conserved mutations, a full stop (.) indicating semiconservative mutations and an empty space ( ) indicating non-conservative mutations.

Along the conserved regions of the various sequences, a section with high conservation was chosen (Figure 7.5) for the placement of both the forward and the reverse primers. The placement was done in such a way that the largest portion of the gene was duplicated and amplified during PCR. Once the forward and reverse

101 | P a g e primers were selected they were blasted in NCBI Primer Blast (Ye et al., 2012) to confirm specificity, GC content and Tm.

7.1.5 General PCR – Primer specificity Following the completion of the conversion of RNA to cDNA, general PCR was employed with the use of specifically designed primers. Primers received from Sigma-Aldrich were at a concentration of 100 μM and were concentrated to 10 μM, the concentration required for the PCR protocol (TaKaRa PCR amplification kit). The reagents used for PCR were Ex Taq (TaKaRa, KA6701BA), 10xEx Taq buffer (TaKaRa, AB1601A), 2.5 mM dNTP (TaKaRa, BK1501A), the ordered forward and reverse primers, double distilled water (DDW) and the cDNA template. A master mix was made with EX Taq (0.1 μL), 10xEx Taq buffer (2 μL), dNTP (1.6 μL) and DDW (13.3 μL) and 17 μL of this master mix was added to each tube. The amount of master mix made depends on the amount of samples you wish to run per PCR procedure, but the ratios of the components in the master mix, as well as the volume of master mix added to the polypropylene PCR tube strip (strip containing 8 tubes), always remained constant. To the master mix in each well 1 μL of the desired forward and reverse primers was added. Finally 1 μL cDNA template was added to each well and the lid was clipped tightly closed. The final mixture in the PCR tubes was placed into the Bio-Rad iCycler PCR System (Bio-Rad Laboratories) at 94˚C for 4 mins followed by 35 cycles of the following protocol: 94˚C for 30 secs, 55˚C for 30 secs, 72˚C for 1 min, 72˚C for 4 mins. After the 35 cycles were completed the iCycler was set to finish on a temperature of 4˚C for ∞, so that the PCR template was maintained at a stable temperature until the products could be further analysed. The 55˚C used for 30 secs during the protocol was the pre-set temperature in this protocol, but this temperature was changes depending on the required Tm of the designed forward and reverse primers. The Tm was given in the information leaflet supplied by Sigma-Aldrich with each set of primers. The temperature setting during this phase of the protocol was never lower than 55˚C or higher than 60˚C.

Amplification products were checked using a 1% agarose gel following the gel electrophoresis procedure.

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7.1.6 Sub cloning Petri dish preparation To 100 mL of the pre-made LB medium, 1.5 g bacteriological agar (Sigma-Aldrich) was added. This mixture was autoclaved for 20 min on the liquid cycle and once cool, 20 μL ampicillin (Sigma-Aldrich) was added. Sterile polystyrene Petri dishes (Sigma-Aldrich) were placed into the fume cupboard and 7 mL of the ampicillin positive LB and agar mixture was added to each Petri dish and allowed to set. Once set, the petri dishes were stored at 4˚C until use.

TOPO® Cloning reaction Cloning of the Taq-Polymerase-amplified PCR products took place with the use of the pCR™2.1 TOPO® vector. Two microliters of the PCR product, 0.5 μL pCR™2.1-

TOPO® vector, 0.5 μL salt solution (1.2 M NaCl and 0.06 M MgCl2) and 2 μL DDW were added to a Corning® Thermowell PCR 8 well strip tubes with caps (one sample per well). The samples were mixed together carefully with a toothpick and incubated at room temperature for 5 mins. Two microliters of this mixture was pipetted into a separate Eppendorf tube per sample.

Rapid One Shot® chemical transformation protocol To each Eppendorf tube, 10 μL of DH5α competent cells (E. coli) was added and the mixture was incubated on ice for 30 mins. The mixture was then placed into a heat block at 42˚C for 30 secs, and placed directly back onto ice for another 5 mins. To each Eppendorf tube, 150 μL of pre-made LB medium was added and the final mixture was pre-incubated on a shaker (Recipro Shaker NR-1, Taitec) at 37˚C for 1 hour.

Pre-prepared petri dishes were removed from the refrigerator and warmed to 37˚C in an incubator. Ten microliters X-gal (Sigma-Aldrich) was pipetted onto the petri dishes and spread evenly over the surface with a flame-sterilised Drigalski spatula. Once dry, 15 μL ampicillin (Sigma-Aldrich) was spread over the surface of the petri dish and allowed to dry. Finally 50 μL of the transformed sample was pipetted into the petri dish and spread evenly over the surface. The Petri dished were then incubated at 37˚C for 16 hours to allow for culture growth.

Analysis of transformants by PCR After the 16 hour incubation period, cultures where checked for any growth. Colonies that appeared white on the growth plate were positive for recombinant DNA. Two to

103 | P a g e six white colonies were gently removed from the plate with a toothpick and mixed into 100 μL LB solution in an Eppendorf tube. This made up the colony solution.

A master mix (MM) was made containing 0.1 μL Ex Taq, 2 μL 10x Ex Taq buffer, 1.6 μL dNTP, 1 μL of the colony solution (LB medium + colony) and 13.3 μL DDW. The volume of the master mix changed according to the number of samples, but the ratios remained the same. To the Corning® Thermowell PCR 8 well strip tubes with caps, 18 μL of the master mix was added to each well. To the 1st, 3rd, 5th and 7th well, 1 μL of the M13 forward and 1 μL M13 reverse primer was added. To the remaining PCR wells the target gene (TG) forward (1 μL) and reverse (1 μL) primers were added.

The same PCR programme and procedure was used as previously described in the General PCR section. The PCR solution was checked for successful amplification through gel electrophoresis. Successful gel electrophoresis resulted in further incubation of the colonies. The remaining 98 μL of the colony solution was pipetted into a 15 mL Falcon tube containing 4 mL LB medium and 4 μL ampicillin. This mixture was placed onto a shaker and incubated at 37˚C for a further 16 hours to increase the colony growth. Successful growth in the LB medium resulted in a milky colour in the LB medium (Figure 7.6B).

Figure 7. 6: Figure showing the both unsuccessful (A) colony culture in LB medium as well as successful (B) colony growth in LB medium after 16 hour incubation period.

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Plasmid purification Following the protocol outlined in the LaboPass™Plasmid Minikit manual, 2 mL LB culture medium was pipetted into a new Eppendorf tube and centrifuged at 11 000 g. for 5 mins to pellet the bacteria. As much supernatant as possible was discarded and the pellet was re-suspended thoroughly by vortexing with 250 μL of Buffer S1. To this mixture 250 μL of Buffer S2 was added and mixed by inverting the Eppendorf tube 4 times; this stage was not vortexed. Three hundred and fifty microliters of Buffer S3 was added and mixed by inverting the Eppendorf tube 4 to 6 times; again, this stage was not vortexed. The mixture was centrifuged at 11 000 g. for 10 mins. The supernatant was transferred to a new spin column and centrifuged at 11 000 g. for 1 min. The flow through was discarded and 300 μL of Buffer PW was added and centrifuged at 11 000 g for 1 min. The flow through was again discarded and centrifuged at 11 000 g for another min. The Eppendorf tube and flow through were discarded and the column was placed into a new Eppendorf tube. Fifty millilitres of EB was added directly to the column filter and centrifuged at 14 000 g. for 1 min. The column was discarded and the resulting flow through was stored at -20˚C. The quantity and quality of the resulting gDNA was checked using a Nanodrop® ND- 1000 Spectrophotometer (Thermo Fisher Scientific).

7.1.6 qRT - PCR The dilution method In order to establish the efficiency of the sequenced primers, the dilution method was used (Pfaffl, 2006). This method requires the serial dilution of the cDNA template and in this case, 20 μL of the original cDNA template was pipetted into an Eppendorf tube and to this 80 μL DDW was added. This was dilution 1. From here a serial dilution was made by adding 2 μL of dilution 1 to a new Eppendorf tube and adding 6 μL DDW to form dilution 2. This procedure was repeated until a total of 6 dilutions were made. A master mix (MM) containing 5 μL SYBR® Green, 0.4 μL sequenced forward, 0.4 μL sequenced reverse primer and 4.2 μL DDW was made. Two microliters of each dilution was pipetted in duplicate into a PCR Multiwell plate (Sigma-Aldrich). To each well 8 μL of the SYBR® Green master mix was added.

The pre-set protocol on the Applied Biosystems 7500/7500 Fast Real-Time PCR System was used. The protocol began by holding the samples at 95˚C for 20 secs. This initial holding period was followed by 95˚C for 30 secs and 60˚C for 30 secs for

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40 cycles. These cycles were then followed by the melting curve cycle at 95˚C for 15 secs, 60˚C for 60 secs and 95˚C for 15 secs.

This protocol was followed in order to establish the efficiency of the sequenced primers.

Primer efficiency calculations The efficiency of the sequenced primers within each of the extracted tissues was calculated based on a linear regression slope of the dilutions of each row (Pfaffl, 2006). The efficiency of each of the sequenced primers in each tissue was calculated using the following equation:

퐸 = 10[−1⁄ 푠푙표푝푒] (Rasmussen, 2001)

The resulting efficiency when using the dilution method ranges from 1.60 to 2 and the closer the value is to 2, the more efficient the primer is (Souazé et al., 1996).

7.2 Results and discussion Primers initially designed for β-actin proved to be unsuccessful and therefore the forward and reverse primers for Silurius meridionalis (Chinese large-mouth catfish) were applied to the tissue samples (Huang et al., 2011). The primers designed through the alignment of homologous genes for transferrin and ER proved to be successful. Homologous genes from fish species such as Ictalurus punctatus (Channel catfish), Pelteobagrus fulvidraco (Yellow catfish), Oreochromis niloticus (Nile tilapia), Oncorhynchus mykiss (Rainbow trout) and Oryzias latipes (Japanese rice fish), were used during the alignment stage and primers were selected within the conserved regions of each gene. Sequenced genes from the order Siluriformes were used wherever possible, but since sequenced genes from this order are scarce, fish species from other orders were used as well. All forward and reverse primers used for each of the sequenced genes are listed in Table 7.1.

Table 7. 1: Forward and reverse primer sets compiled for the amplification of β-actin, Transferrin and Oestrogen Receptor (ER) genes. Gene Forward (5’-3’) Reverse (5’-3’) Beta Actin CCATCTCCTGCTCGAAGTCC GCCCATCTACGAGGGTTACG (β-actin) Transferrin GCTACTACGCTGTAGCTGTAG CCGCGTCCTCCAGAGGTTTC Oestrogen Receptor GATCAGTTAATCATCCTGG CGCACACCTCGCGACGACTTTC (ER)

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Agarose gels run after PCR showed distinct bands of gene amplification in the PCR products for each primer set of each gene. Figure 7.7A shows the bands formed from the 100 bp standard added to each gel for base pair comparison of amplified genes. The bands that follow show amplification products of the primer pairs for ER (1B-4B) and β-actin (1C-4C) and these genes are clearly amplified in all four of the sampled tissues. Figure 7.8 however, shows clear amplification bands for transferrin in liver (2A) and kidney (3A) tissue, but in muscle (1A) and gonadal (4A) tissue the bands appear at the incorrect base pair positions, suggesting that a different gene is amplified in this case. Reasons for the differences in position can be linked to the types of tissues amplified and to the fact that transferrin is a gene related to the binding and transport of iron and so may not be present in muscle and gonadal tissues since these organs are less involved in these physiological functions (Crichton and Charloteaux-wauters, 1987).

Figure 7. 7: Amplification of PCR products for ER (B) and β-actin (C) in muscle (1), liver (2), kidney (3) and gonadal (4) tissue of Synodontis zambezensis.

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Figure 7. 8: Amplification of PCR products for the transferrin gene in muscle (1), liver (2), kidney (3) and gonadal (4) tissue of Synodontis zambezensis.

Successfully amplified genes with satisfactory bands at the correct base pair positions were sent to the Food Assessment and Management Centre (FASMAC. Co. Ltd, Japan) for sequencing. The gene sections sequenced were as follows:

β-actin: TCCATCTCCTGCTCGAAGTCCAGAGCAACATAGCACAGCTTCTCCTTGATGTCA CGCACGATCTCACGCTCGGCCGTGGTGGTGAAGCTGTAGCCACGCTCGGTCA GGATCTTCATCAGGTAGTCCGTGAGGTCACGGCCAGCCAGGTCCAGACGCAG GATGGCATGGGGCAGAGCGTAACCCTCGTAGATGGGC Transferrin: GCTACTACGCTGTAGCTGTAGTAAAGAAAGGCACTAACTTTGGCTTCAAGGACC TTCGTGGGAAGAAGTCCTGTCACACTGGTTTGGGGAAAACTGCAGGCTGGAAC ATCCCCATTGGCACTCTCCTCTCAAAAAAACAAATTCAATGGGGAGGAATCGAT GAGAAACCTCTGGAGGACGCGG ER: GATCAGTTAATCATCCTGGAAAGCTCATCTTCTCCCCAGATCTTGTCCTCAGCA GGGATGAGGGCAGCTGTGTGCAGGGACCAGACCCGGTAGAGTTTCTACCCGG TCTCGAAAGTCGTCGCGAGGTGTGCG These sequences were blasted using the Basic Local Alignment Search Tool (BLAST) (Zhang et al., 2000). The sequence for β-actin was confirmed to be aligned with the β-actin gene for Plecoglossus altivelis (Sequence ID: AB020884.1), Leporacanthicus triactis (Sequence ID: KY945482.1), Leporacanthicus heterodon (Sequence ID: KY945477.1) and Ictalurus punctatus (Sequence ID: XM_017457500.1) with an alignment score of ≥ 200 bp. Confirmation of the sequenced transferrin gene was made between Ictalurus punctatus (Sequence ID: NM_001200320.1), Pelteobagrus fulvidraco (Sequence ID: HM357866.1) and Labrus bergylta (Sequence ID: XM_020643712.1) with an alignment score of ≥ 200 bp. The

108 | P a g e sequenced ER gene was confirmed with an alignment score of 80-200 bp to be aligned with the ER of Ictalurus punctatus (Sequence ID: XM_017456575.1), Pygocentrus nattereri (Sequence ID: XM_017684025.1) and Oryzias melastigma (Sequence ID: XM_024266378.1).These are positive results, however many of the matches were partially matched due to the fact that the entire gene was not sequenced.

7.3.1 Primer efficiency Following the successful sequencing of the above mentioned genes, primers for the sequenced genes were designed. In order to check the efficiency of these subsequently designed primers qRT-PCR was used following the dilution method. The primers created for both ER and transferrin proved inefficient for all of the extracted tissues. The slope of the regression line for ER in liver tissue was -0.10 resulting in an efficiency of 6366805292.62. The ER in kidney tissue, testes, and gonads had slopes of -0.37, -0.18 and -0.23 respectively. These slopes led to efficiencies of 470.29 (kidney), 443126.16 (testes) and 21986.82 (gonads) (Table 7.2). Based on the efficiency calculation, these values are far too high and indicate that amplification and duplication of the genes is not efficient when using these primers (Rasmussen, 2001).

Table 7. 2: Slope of regression lines and resulting efficiencies of ER, transferrin and β-actin forward and reverse primers for successfully sequenced corresponding genes in Synodontis zambezensis. ER Transferrin β-Actin Tissue Slope Efficiency Slope Efficiency Slope Efficiency Liver -0.10 6366805292.62 -0.90 12.86 -3.1 2.10 Kidney -0.37 470.29 -0.45 169.50 -3.54 1.92 Testes -0.18 443126.16 -0.19 183298.07 -3.79 1.84 Gonads -0.23 21986.82 -0.70 26.71 -3.8 1.83

When looking at the slopes for the β-Actin forward and reverse primers, the resulting efficiencies are 2.10, 1.92, 1.84 and 1.83 for liver, kidney, testes and gonads respectively. These results indicate that the primers selected for the sequenced β- actin gene are amplifying and replicating the target gene successfully and efficiently and can therefore be used as a HK gene for further quantitation of gene amplification (Table 7.2).

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Figure 7. 9: Regression lines indicating the efficiency of the β-actin forward and reverse primers within the extracted tissues (liver, kidney, testes and gonads) of Synodontis zambezensis.

The required negative slope (Rasmussen, 2001) can clearly be seen in Figure 7.9, and further indicates the efficiency of the β-actin primers selected for the amplification of the sequenced β-actin gene.

7.4 Conclusions and recommendations Isolation of RNA was successfully completed in all four of the target tissues, and this was evident in the visible 18s and 28s bands achieved during gel electrophoresis. The quality and quantity of the RNA isolated was also satisfactory with the 260/280 readings falling between 1.8 and 2 for good quality samples. Conversion to cDNA as well as identification of successful primers for ER, transferrin and β-actin was successfully completed. All of the tissue samples have been successfully processed up to the point that they can all be tested for new sets of primers so that the final aims of primer efficiency testing and quantitative analysis of gene expression can be completed. The final aim of determining the inhibition and/or excitation of certain genes as a result of exposure to 4,4’-DDT was unfortunately not successfully completed.

The identification of an efficient primer pair for the HK gene β-actin is a very important step towards completing this study. Identification of efficient primers which

110 | P a g e actually amplify the gene of interest can take months to complete, meaning the identification of the HK gene is still an important milestone in this study.

All techniques learnt and applied during this section of the study were completed in Japan at Hokkaido University. These techniques were the objectives used to complete the aims set out for this section of the study. Unfortunately the trip was only three months in length and so the amount of time required to complete all of the analysis was very limited.

It is recommended sequencing of the specified genes be completed and that primer efficiency testing be completed on each of the primers designed from sequenced genes. This should be done in order to confirm the efficiency of each primer pair in the tissues of Synodontis zambezensis and to establish whether more efficient primers need to be designed before qRT-PCR analysis can be completed. This will allow for comparative amplification studies to be completed. Other important sex linked genes will be focussed on in further studies to allow for a more holistic understanding of the endocrine disruption that may be taking place within this species.

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Chapter 8: LC50 exposures and species sensitivity distributions

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8. Introduction Acute toxicity testing of new pesticides in a controlled laboratory environment is a requirement in many countries to determine any future damage that may be caused by the use of unknown compounds (Scholz et al., 2013). These tests are often considered the first stage of chemical toxicity testing, followed by a myriad of other toxicological tests to establish acceptable maximum concentrations of specific chemicals (Chapman, 1995; Kimball and Levin, 1985; Wepener and Chapman, 2012). These tests are used to manage important environmental resources, such as water and ecologically important fish species worldwide (Wepener and Chapman, 2012).

Lowest concentration (LC) or lowest dose (LD) values are calculated with the use of probability unit (Probit) analysis (Finney, 1947). Studies conducted by Chester Bliss in the early 1930s focused on the relationship between the percentage of insecticides and the resulting response in percentage mortality (Bliss, 1939, 1935; Bliss and Stevens, 1937). The graphs created were sigmoidal in nature and were transformed with the use of Probits to represent percentage mortality, into linear graphs in order for the completion of regression analysis. A table of Probits was compiled by DJ Finney in 1947, and this table (Table 8.1 - Statistics) forms the basis of Probit analysis used in this study (Finney, 1947). The transformation of mortality into Probits and the log10 of the dose results in the linearity of the curve (Finney and Stevens, 1948). The resulting linear equation is used for the calculation of the concentration that will cause mortality of 50% of the individuals, the LC50 (Finney, 1947; Finney and Stevens, 1948; Hayes and Kruger, 2014).

Species sensitivity distributions (SSDs) are often incorporated into ecological risk assessments in order to give a more holistic view of the toxic effects a specific chemical or mixture of chemicals, may have on the environment (Solomon et al., 1996; Steen et al., 1999). The benefit of using a variety of organisms to assess the toxicity of a specific chemical is that there is greater statistical confidence offered when calculating the toxicity limits of a system (Wheeler et al., 2002).

Three fish species were used for the assessment of LC50 values in response to exposure to 4,4’-DDT. The reason for using three different fish species is to attempt to determine the sensitivity of each fish in relation to the other and make

113 | P a g e comparisons based on previous studies done on each species (Wheeler et al., 2002). Given their broad tolerances to water quality and temperature, these three fish species, Oreochromis mossambicus, Poecillia reticulata and Synodontis zambezensis may potentially occur within the same freshwater aquatic systems (Bills et al., 2010; Deacon et al., 2015; Smit et al., 2016).

Oreochromis mossambicus is a tilapia native to South Africa (Blue kurper), popular in aquaculture and economically important (Cambray and Swarts, 2007). This fish species is robust and adapts well to changing environmental conditions, which has allowed it to become a very successful invasive species in Hawaii and California (Courtenay and Robins, 1989) . They have also become invasive in Australia, mostly due to their ability to adapt to wide ranges of salinity (euryhaline organisms) (Russell et al., 2012).

Figure 8. 1: Oreochromis mossambicus, picture taken from Cambray and Swarts (2007). Photo credit Dr D. Tweddle.

This species was one of 50 fish species intentionally introduced into the inland waters of California for the control of mosquito populations (Moyle, 1976), but its robustness and general adaptability led to it outcompeting many native fish populations (Moyle, 1976). They have been extensively used in testing the effects of metals (Basha and Rani, 2003; Van Dyk et al., 2007), organophosphate pesticides (Rao et al., 2003) and health assessments of South African rivers (Marchand et al., 2012; Smit et al., 2016). There have, however, been no previous assessments of

LC50 values for 4,4’-DDT using O. mossambicus as the test species.

Poecilia reticulata or the guppy is native to Brazil, Barbados and Jamaica (Araújo et al., 2009; Casatti et al., 2009), and an invasive species in South African waters (Tavakol et al., 2017). This species is also highly adaptable allowing them to survive

114 | P a g e and flourish in a variety of habitats (Magurran, 2005). Their introduction into African freshwater systems relates back to their use to control mosquito populations responsible for the spread of malaria (Magurran and Phillip, 2001). They feed on a variety of aquatic insect larvae, including mosquito larvae, which led to their use as an organic control measure in malaria stricken countries (Dussault and Kramer, 1981).

Figure 8. 2: A male (above) and female (below) Trinidadian guppy (Deacon et al., 2015).

Guppies have been used in studies of ecology, evolution and behavioural adaptations (Magurran, 2005). They have also been used in toxicity testing of a variety of chemicals including deltamethrin (Viran et al., 2003), β-cypermethrin (Polat et al., 2002) and α-cypermethrin (Yılmaz et al., 2004). They have also been used as test organisms in studies focused on the effects of chemical mixes (Hermens and Leeuwangh, 1982) and DDT and its degradation products (Kristensen et al., 2006; Morgan, 1972; Weis, 1974) in the environment.

The focus species of this study, Synodontis zambezensis has been discussed in detail in chapter 3 and 4 of this thesis. Their use as an indicator species occurred fairly recently (Smit et al., 2016), and they are not well-known as a biological indicator species.

The aim of this section of the study is to show that Synodontis zambezensis can be used successfully in acute toxicity studies. They are easy to breed in captivity and house in a laboratory environment (Chapter 4), and the results of this section of the study may show that they are more sensitive to 4,4’-DDT than the other well-known

115 | P a g e biological indicator species, O. mossambicus and P. reticulata. The aims of this section of the study will be completed through the methods and procedures outlined in the materials and methods section.

8.1 Materials and methods 8.1.1 Fish species

All three fish species used during the juvenile LC50 exposure experiments were bred in the aquarium at the University of Johannesburg. The breeding procedures used for Synodontis zambezensis have been outlined in Chapter 4 in the published article “Artificial breeding and embryonic development of Synodontis zambezensis (Peters, 1852)”. The procedures followed for the breeding of the other two fish species are standardised procedures used in the University of Johannesburg aquarium (Oliveira and Almada, 1998; Snyder et al., 1996). Fish were housed according to the guidelines outlined in the SANS document 10386 (SANS, 2008).

8.1.2 Acclimation Fish were acclimated in an environmental room used for the 96-hour exposure experiment for one week before exposure experiments commenced. The environmental room was kept at a constant temperature of 25˚C with a 12h/12 hour day/night photoperiod.

8.1.3 Experimental setup Prior to testing, a stock solution made up of 1 L 95% ethanol (Merck Millipore) and a 100 mg vial of supplied 4,4’-DDT (Sigma-Aldrich, South Africa) was prepared. The 95% ethanol was used as the organic solvent for the 4,4’-DDT since its solubility in water is so low (0.12 ppm at 25˚C) it is considered insoluble (NIST, 2015) and therefore requires dissolution in an organic solvent before addition to any water source (USEPA, 1979). The final nominal concentration of the stock solution used was 98 700 ng/mL or 98.7ppb. The water used for the exposure was reconstituted water or fish medium (IWQS, 1998) consisting of 50 L deionized water to which:

23.52 g calcium chloride (CaCl2), 9.86 g magnesium sulphate (MgSO4), 5.18 g sodium bicarbonate (NaHCO3), 0.46 g potassium chloride (KCl) was added (IWQS, 1998).

The exposure concentrations used were based on the LC50 values for three fish species; Oncorhynchus mykiss (Rainbow Trout), Pimephales promelas (Fathead Minnow) and Lepomis macrochirus (Bluegill) which range from 3.4 ng/mL to 10

116 | P a g e ng/mL (Sigma-Aldrich, 2014). The LC50 values are listed as part of the Material Safety Data Sheet (MSDS) of the 4,4’-DDT used during the exposure (Sigma- Aldrich, South Africa). The test conditions were static and the environmental room was maintained at a constant temperature of 25˚C with a 12h/12 hour day/night photoperiod. Five hundred millilitre glass beakers filled to a volume of 400 mL were used for the 96-hour exposure. Each concentration was tested in duplicate with 5 fish (14-day old juveniles) of each species added to their own set of beakers (Figure 8.3). The beakers were labelled Control, Solvent control and concentrations (0.6 ng/mL, 1.2 ng/mL, 2.5 ng/mL, 3.7 ng/mL, 5.6 ng/mL, 7.4 ng/mL, 8.6 ng/mL, 10.5 ng/mL, 12.3 ng/mL, 13.6 ng/mL and 15.4 ng/mL). Sub-lethal and lethal effects were assessed every 24 hours and compared with the control groups.

Figure 8. 3: Set up of 500 mL beakers. Both controls and all exposure concentrations ranging from 0.6 ng/mL to 15.4 ng/mL were set up in duplicates in an environmental room at 25˚C with a 12h/12h day/night photoperiod.

Comparisons of exposure concentration groups are made with the control group in order to determine whether the test meets the “pass” requirements of a <10% mortality in all controls. Fish were considered to be dead when no observable movement was present and if they did not respond to touch at all. Any observable changes in behaviour were noted.

Testing was conducted in duplicate because the availability of the test organisms in the required age group restricted the number of possible replicates.

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8.1.4 Water quality

Standard in situ water quality parameters, oxygen percentage (O2%), electrical conductivity (EC), temperature (˚C) and pH, were measured in each tank at 24-hour intervals. The first readings were taken 20 mins after the addition of the relevant concentrations to allow for a homogenous mixture of water and 4,4’-DDT stock solution or 95% ethanol in the case of the solvent control. The water parameters were measured using a Eutech Multi-Parameter (PCTestr™ 35) instrument.

8.1.5 Statistics All water quality data was tested for homogeneity using a Chi-square test (Hayes and Kruger, 2014). This data was subsequently log transformed and SPSS version 25 (IBM Software Group) was used for one-way analysis of variance (ANOVA). Significant differences (p≤0.05) between concentrations and skewness of data were determined using the Tukey Post Hoc test (Yohannes et al., 2013).

Transformative Probit analysis was applied to the data recorded during the 96-hour exposure of the three fish species to 4,4’-DDT at varying concentrations. The exposure concentrations in ng/mL and the mortalities recorded at each concentration were input into a Probit analysis spreadsheet (Finney, 1947; Finney and Stevens,

1948). The resulting graphs, allowed for the calculation of the LC50 value for each of the tested fish species and comparisons were made between their sensitivities in relation to one another.

The steps that were followed during Probit analysis were as follows:

The exposure doses used (λ) were converted to log10 concentrations (Equation 1) and these log10 values become x, and are plotted on the x-axis.

푥 = 푙표푔10휆 Equation 1 (Finney, 1947)

The empirical response rate, p, was calculated based on Equation 2.

푟 푝 = ⁄푛 Equation 2 (Finney and Stevens, 1948)

Where r equals the total number of individuals that respond (in this case death) at a specific concentration and n equals the number of individuals exposed at a specific concentration (in this case 10 per concentration). The empirical responses (p) were converted to empirical probits (y) based on the Table 8.1 (Finney, 1947).

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A dose response curve was drawn using x (Equation 1) the empirical probits (y) calculated using equation 2, and the regression equation was derived from the resulting graph. Empirical probits resulting from the percentage mortality that were less than 1 or greater than 7 were not taken into account when graphing the response curve. These values have little to no effect on the estimation of the LC50 (Hayes and Kruger, 2014).

Table 8. 1: Table showing the transformation values of percentage mortality into probability units (Probits) (Finney, 1947).

Transformation of percentages to Probits % 0 1 2 3 4 5 6 7 8 9 0 - 2.67 2.95 3.12 3.25 3.36 3.45 3.52 3.59 3.68 10 3.72 3.77 3.82 3.87 3.92 3.96 4.01 4.05 4.08 4.12 20 4.16 4.19 4.23 4.26 4.29 4.33 4.36 4.39 4.42 4.45 30 4.48 4.50 4.53 4.56 4.59 4.61 4.64 4.67 4.69 4.72 40 4.75 4.77 4.80 4.82 4.85 4.87 4.90 4.92 4.95 4.97 50 5.00 5.03 5.05 5.08 5.10 5.13 5.15 5.18 5.20 5.23 60 5.25 5.28 5.31 5.33 5.36 5.39 5.41 5.44 5.47 5.50 70 5.52 5.55 5.58 5.61 5.64 5.67 5.71 5.74 5.77 5.81 80 5.84 5.88 5.92 5.95 5.99 6.04 6.08 6.13 6.18 6.23 90 6.28 6.34 6.41 6.48 6.55 6.64 6.75 6.88 7.05 7.33 0-0 0-1 0-2 0-3 0-4 0-5 0-6 0-7 0-8 0-9 99 7.33 7.37 7.41 7.46 7.51 7.58 7.65 7.75 7.88 8.09

The regression equation:

푦 = 푚푥 + 푐 Equation 3 was derived from each regression analysis to find the LC50 value for each species, the empirical probit 5.00 was substituted into the equation for x.

The equation was reshuffled in order to solve for y as follows:

푦 = (5 − 푐)⁄ 푠푙표푝푒 Equation 4

The resulting y value gave the log10 concentration at which 50% of the individuals would die. The antilog of this value was computed, resulting in a concentration

(ng/mL) corresponding to the required LC50 value.

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8.2 Results 8.2.1 Water quality The water quality parameters did not change much over the 96-hour exposure period

(Table 8.2). The main change noted was a sudden drop in oxygen percentage (O2%) just after the addition of the various concentrations and the ethanol in the solvent control. The pH ranged from 7.7 to 8.3, which is within the pH range tolerated by all three species (Bills et al., 2010; Magurran, 2005; Skelton, 2001). The EC ranged from 1103 µS/cm to 1147 µS/cm with the higher EC readings noted in the beakers containing the higher concentrations of 4,4’-DDT.

Table 8. 2: In situ water quality parameters (mean ± SD) and range readings taken of the exposure beakers during the course of the 96-hour 4,4’-DDT LC50 exposure experiment. electrical conductivity (EC), pH, oxygen percentage (O2%) and temperature (˚C). Conc. EC (µS/cm) Temp. (˚C) pH O (%) (ng/mL) 2 1126.80 ± 0.04 23.56 ± 0.05 8 ± 0.06 75.85 ± 8.13 Control (1103 - 1147) (23.40 - 23.70) (7.8 - 8.2) (48.2 - 94.5) Solvent 1124.4 ± 5.25 23.68 ± 0.07 8.02 ± 0.07 73.36 ± 7.88 control (1108 - 1143) (23.5 - 23.9) (7.9 - 8.3) (46.6 - 93.4) 1125.6 ± 3.74 23.64 ± 0.06 7.82 ± 0.07 71.34 ± 5.59 0.60 (1116 - 1138) (23.4 - 23.8) (7.7 - 8.1) (53.1 - 82.3) 1120 ± 3.75 23.66 ± 0.04 7.94 ± 0.08 68.94 ± 6.00 1.20 (1107 - 1131) (23.6 - 23.8) (7.8 - 8.3) (47.4 - 82.6) 1123.4 ± 3.04 23.76 ± 0.06 7.8 ± 0.07 75.18 ± 8.02 2.50 (1116 - 1134) (23.6 - 24) (7.7 - 8.1) (48.2 - 94.6) 1121.6 ± 3.48 23.78 ± 0.06 7.9 ± 0.09 75.64 ± 9.48 3.70 (1111 - 1133) (23.6 - 24) (7.7 - 8.3) (46.6 - 108.2) 1122.8 ± 4.57 23.82 ± 0.04 7.8 ± 0.07 69.28 ± 4.88 5.60 (1106 - 1135) (23.7 - 24) (7.6 - 8.1) (53.1 - 69.3) 1121.6 ± 3.29 23.8 ± 0.08 7.96 ± 0.10 69.72 ± 6.58 7.40 (1111 - 1132) (23.5 - 24) (7.8 - 8.4) (47.4 - 89.8) 1125.2 ± 4.20 23.88 ± 0.03 7.78 ± 0.09 75.78 ± 3.94 8.60 (1113 - 1138) (23.8 - 24) (7.7 - 8.1) (59.2 - 83.3) 1122.4 ± 4.53 23.9 ± 0.04 7.92 ± 0.09 75.68 ± 4.49 10.50 (1109 - 1136) (23.8 - 24) (7.7 - 8.3) (58.4 - 89.2) 1126.8 ± 5.50 23.9 ± 0.03 7.72 ± 0.08 75.78 ± 3.94 12.30 (1113 - 1147) (23.8 - 24) (7.7 - 8) (59.2 - 83.3) The range of temperatures found during the exposure experiment were similar, but still fell within the temperature tolerance range of all three species (Bills et al., 2010; Magurran, 2005; Skelton, 2001), ranging from 23.4˚C to 24˚C. Mean oxygen percentage is required to remain above 60% throughout the exposure period

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(OECD, 1992), and this was achieved during the exposure period with oxygen percentages ranging from a mean of 68.94% to 75.85%. There were no significant differences (p≤0.05) found between the different readings throughout the exposure concentrations over the 96-hour experimental period.

8.2.2 LC50 Transformative Probit analysis After application of transformative Probit analysis, regression curves for each species were drawn separately. The resulting regression equations were then used to determine the LC50 values for each species (Figure 8.4A, B and C). The regression curves show the concentrations at which deaths occurred and do not include the control or solvent control as there were no deaths in any of these beakers for any of the three species. Figure 8.4D shows the species sensitivity distribution of the three species in relation to one another.

In Figure 8.4A the regression analysis for S. zambezensis is illustrated. Mortality within this species during the 96-hour exposure period began in the lowest exposure concentration of 0.6 ng/mL at 20% mortality. This percentage increased as the exposure concentrations increased resulting in a regression equation of y = 1.8332x + 4.4709 and an R² = 0.9126. The R2 value is very close to 1, meaning that almost 100 % of the variation in the data is explained by the regression equation. Mortality in this species reached 100% in exposure concentrations 8.6 ng/mL, 10.5 ng/mL and 12.3 ng/mL. Mortality occurred in 71.11 % of the total S. zambezensis juveniles (110) exposed during this experiment.

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Figure 8. 4: Linear regression graphs from 96-hour exposure to 4,4’-DDT of juvenile S. zambezensis (A), P. reticulata (B) and O. mossambicus (C) as well as the distribution of sensitivity between these three species (D).

The linear regression of Poecilia reticulata in Figure 8.4B shows the resulting linear regression equation of y = 2.9659x + 2.1649 and the R² = 0.8142. The lower R2 value means that less of the variation in the data is explained by the linear regression, only about 80%. There is no mortality seen in this species until the third exposure concentration of 2.5 ng/mL and after 96-hours, the highest percentage mortality is only 80% of the exposed organisms in the highest exposure concentration. Only 27.78% of the total exposed organisms (110) died during the 96- hour exposure period.

The regression analysis for Oreochromis mossambicus (Figure 8.4C) resulted in a regression equation of y = 2.0233x + 3.9618 and an R² = 0.6066. The R2 value is much lower than those found in either of the other species and only about 60% of the variation in the data is explained by this regression analysis. Mortality occurred at 20% in the lowest exposure concentration of 0.6 ng/mL, but no mortality at all was seen in the next exposure concentration of 1.2 ng/mL throughout the 96-hour

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exposure period. One hundred percent mortality resulted in this species at the 10.5 ng/mL and 12.3 ng/mL exposure concentrations.

Table 8. 3: Species, regression equation and the solutions for x as described in the statistics section. The resulting LC50 values indicated the sensitivities of each of the exposed species.

Species Regression equation Solve for x Antilog x LC50 S. zambezensis y = 1.8332x + 4.4709 x = (5-4.4709)/1.8332 = 0.288621 Antilog(0.288620991) 1.94 O. mossambicus y = 2.0233x + 3.9618 x = (5-3.9618)/2.0233 = 0.513122 Antilog(0.513122127) 3.26 P. reticulata y = 2.9659x + 2.1649 x = (5-2.1649)/2.9659 = 0.95589871 Antilog(0.955898715) 9.03

The species sensitivity distribution (Figure 8.4D) clearly shows that S. zambezensis is more sensitive than both O. mossambicus and P. reticulata and this is

corroborated by the resulting LC50 values in Table 8.3.

8.3 Discussion 8.3.1 Water quality The values of EC in the exposure beakers were expected to be slightly higher than those that would be found in borehole water, since the exposure medium was reconstituted water and was a mixture of various salts (IWQS, 1998). Since EC relates to the electrical nature of the dissolved salts within a solution, the higher EC values within this experiment are expected (Sawyer and McCarty, 1978). The readings taken of EC did not fluctuate much during the exposure period, meaning the environment remained stable throughout. The reconstituted water used during this experiment was specifically formulated for the use in experimentation involving fish (IWQS, 1998), so the higher than normal EC would not have had any adverse effects on the fish during this experiment.

The initial decrease in oxygen percentage after the addition of the exposure concentrations and ethanol in the case of the solvent control was remedied by the aeration of the fish medium for 10 mins every 24 hours. This resulted in a stability of oxygen percentage throughout the exposure period. The reason for the drop in oxygen percentage after the addition of the exposure concentrations is not known.

8.3.2 LC50 Transformative Probit analysis The resulting linear regression, LC50 values and species sensitivity distribution clearly show that S. zambezensis is more sensitive to the toxic effects of 4,4’-DDT

than O. mossambicus and P. reticulata. The LC50 concentrations for both S.

123 | P a g e zambezensis and O. mossambicus are lower than Oncorhynchus mykiss (Sigma-

Aldrich, 2014), and all three of the species used in this study have lower LC50 concentrations than both Pimephales promelas and Lepomis macrochirus (Sigma- Aldrich, 2014).

8.4 Conclusions and recommendations

The aim of this section of the study was to establish the LC50 values for the three focus fish species and this was done successfully. Transformative Probit analysis was successfully applied, resulting in LC50 values for all three species. The exposure experiments resulted in no mortality in either the control or the solvent control in any of the three species used confirming the validity of the completed experiments (OECD, 2014, 1992).

The confirmation of the LC50 value for S. zambezensis and the comparison of this value with previously established LC50 values of other fish species (Sigma-Aldrich, 2014) shows that S. zambezensis is a sensitive fish species, which can be considered for further laboratory experimentation of toxic chemicals. They have proved to be easy to breed and keep under laboratory conditions (Chapter 4) and their increased sensitivity to chemicals such as 4,4’-DDT make them ideal for testing the effects that toxic chemicals may have on the ecology of many river systems in South Africa.

Their importance as a food source adds to their appeal as a laboratory test organism, owing to the fact that human health risk assessments can be done in communities using this species as a source of protein.

It is recommended that the Fish Embryo Test (FET) (Braunbeck et al., 2015) be completed using the fertilised eggs of S. zambezensis. The inclusion of this test in future studies of S. zambezensis will allow for a better understanding of the toxic effects that chemicals, including 4,4’-DDT, may have on the growth, development and ultimate breeding success of this economically important fish species. Follow-up species sensitivity tests should also be completed to include the sensitivity of invertebrate species such as Daphnia magna as well as algal species Scenedesmus quadricauda (Sigma-Aldrich, 2014).

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Chapter 9: Conclusions and recommendations

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9. Conclusions and recommendations The general effects of 4,4’-DDT have been discussed at length throughout this thesis and establishing its specific effects to Synodontis zambezensis have been presented in the results sections of each chapter. Since the entire study was based around specific 4,4’-DDT concentrations and exposure periods, specific time related conclusions can be drawn and linked directly to the use of 4,4’-DDT in the environment. The continued use of 4,4’-DDT in the environment and its potential to bio-accumulate in S. zambezensis poses a threat to human populations relying on this species as a source of food (Coetzee et al., 2015; Smit et al., 2016). The conclusions and recommendations drawn from the various chapters are therefore, not only important in terms of maintaining wild fish populations, but also in terms of the human communities surrounding river systems that contain this fish species.

9.1 Laboratory based exposures Having been successfully used as a bio-indicator species as part of a large field- based study (Smit et al., 2016), this laboratory-based study allows for clearer understanding of the bio-accumulation potential of S. zambezensis. The results presented in chapter 5, indicate that an acute exposure of 4,4’-DDT to adult S. zambezensis can result in bio-accumulation directly from the aquatic environment. The lowest exposure concentration of 1.97 ng/mL resulted in 76.92% bio- accumulation of 4,4’-DDT equivalents over a 96-hour exposure period. It is therefore resonable to conclude, that should an exposure at this concentration occur in a river system, following the application of 4,4’-DDT to surrounding areas in the form of IRS, wild fish populations have the potential to accumulate just as much if not more 4,4’- DDT, over the same period of time. Considering that 35.5% of the 4,4’-DDT available to fish in the wild is accumulated through the food chain (Macek and Korn, 1970), these levels may increase in wild fish with the inclusion of accumulation through the food chain. The risk posed to communities relying on this species as a source of food is clear, and the results of this study may contribute to decreasing or managing the concentrations of 4,4’-DDT applied during IRS so as to decrease the concentrations that could potentially enter surrounding freshwater systems. Management of the applied concentrations of 4,4’-DDT would also include testing to make sure that the levels that are applied are still effective at controlling the spread of malaria through the control of the Anophlese mosquito.

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It was hypothesised that higher concentrations of 4,4-DDT would result in higher accumulation within the tissue of the exposed adult S. zambezensis. This hypothesis was accepted, since the accumulated concentrations of 4,4’-DDT, 4,4’-DDE and 4,4’- DDD all steadily increased as the exposure concentrations increased. The accumulated percentage equivalents of the initial 4,4’-DDT concentrations did however decrease as the exposure concentrations increased. The decreasing accumulation of the 4,4’-DDT equivalents may be linked to the increasing concentrations of mucus secreted by the exposed fish into the water as explained in the results section of chapter 5.

The highest exposure concentration of 63.17 ng/mL used during the acute exposure of adult S. zambezensis was six times higher than the upper limit of previously calculated LC50 concentrations for Oncorhynchus mykiss (Rainbow Trout), Pimephales promelas (Fathead Minnow) and Lepomis macrochirus (Bluegill) (Sigma-Aldrich, 2014) juveniles, but the concentrations were not expected to cause mortality in the adult S. zambezensis. These concentrations were used both to indicate the sensitivity of S. zambezensis adults when compared to juveniles of the above mentioned species, as well as to determine possible resulting biochemical responses after an acute exposure period.

It was also hypothesised that the exposure of 14-day old S. zambezensis, Oreochromis mossambicus and Poecilia reticulata to varied concentrations of 4,4’-

DDT will allow for the calculation of the LC50 values for all three fish species and subsequent SSDs for each of the three species. This hypothesis was accepted since the resulting mortalities that took place during the 96-hour exposure allowed for the

Probit analysis and final calculation of the LC50 concentrations for all three juvenile fish species. The SSD between these three fish species was also completed and S. zambezensis was by far the most sensitive of the three species used in this study.

Following the exposure of the juvenile S. zambezensis and the final calculation of the

LC50 of 1.94 ng/mL, it is clear that the juvenile S. zambezensis are far more sensitive than adult S. zambezensis, as well as juveniles of Oncorhynchus mykiss, Pimephales promelas, Lepomis macrochirus, Oreochromis mossambicus and Poecilia reticulata. The low sensitivity of S. zambezensis adults demonstrated with the lack of mortality in the high exposure concentrations, indicates that this species may survive in very harsh environments with high concentrations of contaminants,

127 | P a g e provided they have reached the adult stage. However, the high sensitivity of the juvenile S. zambezensis may indicate that this species could suffer a loss of juveniles early on in their development due to increased contamination of river systems and this may lead to an overall decline in the wild population. The ability for the adult S. zambezensis to withstand higher concentration of contaminants may mean that their potential to bio-accumulate pesticides specifically, is even higher than initially expected. The human health risk posed by this species to communities could be even higher than initially assumed (Coetzee et al., 2015; Smit et al., 2016).

The aims and objectives relating to this section of the study were successfully completed and the results have been listed and discussed in detail in the relevant chapters.

9.1.1 Recommendations Further studies recommended include chronic exposures of S. zambezensis to 4,4’- DDT, to allow for the calculation of a species specific Bioconcentration Factor (BCF) (USEPA, 1996). Chronic exposure experiments are required for the calculation of the BCF because the uptake phase is a mandatory 28 days (USEPA, 1996), followed by a depuration period and a final calculation of the ratio between the two. Rerunning of the acute exposure experiments is also recommended at the same exposure concentrations as well as concentrations higher than those used during these experiments for possible determination of the concentration at which 50% mortality

(LC50) is caused in the adult S. zambezensis. The Fish Embryo Test (FET) (Braunbeck et al., 2015) is recommended for future studies since it can be used as an alternative to the Acute Fish Toxicity Test (Braunbeck et al., 2015; OECD, 2014), in the detemination of the LC50 of various species. The inclusion of this test in future studies of S. zambezensis will allow for a better understanding of the toxic effects that chemicals, including 4,4’-DDT, may have on the growth, development and ultimate breeding success of this economically important fish species. Follow up species sensitivity tests should also be completed to include the sensitivity of invertebrate species such as Daphnia magna as well as algal species Scenedesmus quadricauda (Sigma-Aldrich, 2014).

9.2 Biomarker analysis Physiological adaptations in the form of biomarkers often occur bimodally after exposure to toxicants in the environment (Forbes et al., 2006). The use of acute

128 | P a g e toxicity testing in this study resulted in biomarker responses that may not be a true indication of the full adaptation potential of S. zambezensis.

It was hypothesised that the exposure of adult S. zambezensis to increasing concentrations of 4,4’-DDT would result in corroborative increased excitation/inhibition of biochemical responses seen in biomarkers of exposure and effect. This hypothesis was accepted based on the results discussed and the increases in biomarker response, although bimodal in nature, as the exposure concentrations increased.

The results recorded indicated increased oxidative stress in the gills and liver tissue in CAT, GSH and SOD. The fluctuations across the exposure concentrations are a result of the short exposure period used. The physiological responses seen in the liver tissue are less pronounced than those seen in the gills in SOD, and this could be as a result of the gills being in direct contact with the exposure medium, allowing for a more rapid excitation of the SOD antioxidant system. The higher CAT and lower SOD concentrations in the liver tissue is a clear indication of this first line of defence antioxidant system (Van der Oost et al., 2003) working to stabilise the ROSs present within the liver as a result of 4,4’-DDT exposure. The differences seen within the biomarkers of effect indicate that an acute exposure can elicit changes within the ROS and antioxidant homeostasis resulting in predictable future damage to tissues.

The biomarkers of exposure showed very few excitatory or inhibitory responses and were relatively stable throughout the exposure concentrations. The overall energetics also remained stable when looking at total CEA, but there are visible changes in the glucose content when looking specifically at the Ea concentrations. The stability of overall CEA means that although responses in oxidative stress are present, there has been no irreversible damage caused to proteins or lipid containing structures like cell membranes. The damage that may have been caused is visible in the excitation of the oxidative stress biomarker responses, but these changes may have stabilised again should the toxicant have been removed and the fish placed back into freshwater for a further 96-hours.

It was further hypothesised that RNA expression of transferrin and ER genes would increase or decrease in correlation with varied concentrations of 4,4’-DDT following the acute exposure of S. zambezensis. This hypothesis was unfortunately neither

129 | P a g e rejected nor accepted since the fluorescent analysis of gene expression was not completed.

The use of RNA as a biomarker in this study yielded promising results in terms of the identification of a successfully sequenced HK gene, β-actin. The application of fluorescent analysis to determine increases or decreases in gene expression was only applied to determine the efficiency of the successfully sequenced primer pairs for the HK gene and two other possible “biomarker” genes. Unfortunately the sequenced primer pairs for the genes transferrin and oestrogen receptor (ER) were inefficient, but a base has been established for future primer pairs to be ordered an examined.

The aims and objectives relating to this section of the study were successfully completed and the results have been listed and discussed in detail in the relevant chapters.

9.2.1 Recommendations The inclusion of biomarkers which did not form the focus of this study is recommended for future studies. These biomarkers would include heat shock proteins (HSP), blood cortisol and lactate levels and the ratio calculation of GSSG to GSH (Giustarini et al., 2013). Histopathological analysis of the gill, liver and gonadal tissue is also recommended, with specific focus on the gills since they are in direct contact with the exposure medium (Rola et al., 2012). Sub-sampling of fish at 24- hour intervals during a chronic exposure is recommended to include time specific biomarker changes in S. zambezensis.

Primer efficiency of successfully sequenced primer pairs is recommended for future studies. Establishing the efficiency of these primer pairs is very important for the successful completion of this section of the study. It is recommended that further research be done into the Synodontis genus with respect to genetic changes and adaptations. Identifying specifically important genes for S. zambezensis will be essential for the successful understanding of the effects of 4,4’-DDT at a genetic level.

9.3 Artificial breeding of Synodontis zambezensis Artificial breeding and subsequent rearing of S. zambezensis was completed successfully (Volschenk et al., 2018). This is an important step in the continued use

130 | P a g e of this species in laboratory based assessments of toxicants that may enter the environment.

Based on the results of the LC50 and acute exposure experiments, this species has been established as a sensitive alternative for future laboratory exposure experiments. They are easily maintained under laboratory conditions and their medium size allows for the exposure of more individuals per concentration than other larger experimental species such as C. gariepinus. The importance of S. zambezensis as a food source adds to their appeal as a laboratory test organism, owing to the fact that human health risk assessments can be done in communities using this species as a source of protein.

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Appendix A: Raw Data

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Table A: Raw bio-accumulation data from exposure concentration 1.97 ng/mL to exposure concentration 7.90 ng/mL. Concentrations of 4,4’- DDE, 4,4’-DDD and 4,4’-DDT extracted from the Synodontis zambezensis muscle tissue. Both the Wet Weight [WW] and Lipid Weight [LW] concentrations are listed and calculations were done based on the EPA Method 8081B, organochlorine Pesticides by Gas Chromatography.

Lipids Lipids Lipids Conc. Fish Muscle Total Lipid 4,4'- [WW] [LW] 4,4'- [WW] [LW] 4,4'- [WW] [LW] (pre (post in 2ml (ng/mL) # Mass Lipids % DDE ng/g ng/g DDD ng/g ng/g DDT ng/g ng/g weight) weight) (g) 13 7.18 12.21 12.31 0.10 0.50 6.94 0.77 1.08 15.52 2.00 2.79 40.20 1.81 2.52 36.27

14 8.16 12.24 12.39 0.14 0.70 8.58 0.78 0.96 11.18 1.73 2.12 24.72 1.91 2.34 27.26 1.97 15 4.91 12.07 12.19 0.11 0.57 11.56 1.10 2.23 19.31 1.11 2.26 19.55 1.27 2.59 22.39 MEAN 9.00 0.88 1.42 15.34 1.61 2.39 28.16 1.66 2.48 28.64

SD 1.55 0.15 0.57 3.32 0.37 0.29 8.77 0.28 0.10 5.75

SE 0.69 0.07 0.26 1.49 0.17 0.13 3.92 0.13 0.05 2.57

16 6.85 12.32 12.46 0.14 0.69 10.12 0.78 1.14 11.26 1.75 2.56 25.28 0.07 0.10 0.98 17 7.46 12.24 12.36 0.12 0.61 8.18 0.88 1.18 14.40 2.01 2.69 32.93 1.98 2.65 32.39

18 5.09 12.23 12.33 0.10 0.50 9.90 0.78 1.53 15.49 1.75 3.44 34.76 1.92 3.77 38.03 3.95 19 5.54 12.12 12.15 0.03 0.15 2.73 0.83 1.50 55.14 0.78 1.42 51.93 1.03 1.85 67.94 20 3.44 12.18 12.20 0.02 0.12 3.40 0.94 2.73 80.31 0.79 2.30 67.59 0.70 2.05 60.18 MEAN 6.87 0.84 1.62 35.32 1.42 2.48 42.50 1.14 2.08 39.90

SD 3.18 0.06 0.58 27.67 0.52 0.65 15.27 0.73 1.20 23.55

SE 1.01 0.02 0.18 8.75 0.17 0.21 4.83 0.23 0.38 7.45

21 10.22 12.26 12.36 0.10 0.51 5.02 0.77 0.75 14.94 2.21 2.16 43.11 2.09 2.05 40.83 22 5.30 11.96 12.01 0.05 0.23 4.28 0.74 1.39 32.52 1.70 3.21 74.86 1.98 3.73 87.20

23 5.22 12.11 12.17 0.06 0.32 6.21 0.76 1.45 23.39 1.99 3.81 61.37 1.94 3.72 59.89 7.90 24 6.06 13.29 13.34 0.05 0.24 4.01 0.50 0.82 20.50 0.53 0.88 21.94 0.70 1.15 28.66 25 7.92 12.30 12.35 0.06 0.29 3.60 0.37 0.47 13.07 0.49 0.62 17.35 0.63 0.79 22.00 MEAN 4.62 0.63 0.98 20.88 1.39 2.14 43.73 1.47 2.29 47.72

SD 0.92 0.16 0.38 6.90 0.73 1.25 22.14 0.66 1.24 23.58

SE 0.29 0.05 0.12 2.18 0.23 0.40 7.00 0.21 0.39 7.46

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Table A continued: Raw bio-accumulation data from exposure concentration 15.79 ng/mL to exposure concentration 63.17 ng/mL. Concentrations of 4,4’-DDE, 4,4’-DDD and 4,4’-DDT extracted from the Synodontis zambezensis muscle tissue. Both the Wet Weight [WW] and Lipid Weight [LW] concentrations are listed and calculations were done based on the EPA Method 8081B, organochlorine Pesticides by Gas Chromatography.

Lipids Lipids Lipids Conc. Fish Muscle Total Lipid 4,4'- [WW] [LW] 4,4'- [WW] [LW] 4,4'- [WW] [LW] (pre (post in 2ml (ng/mL) # Mass Lipids % DDE ng/g ng/g DDD ng/g ng/g DDT ng/g ng/g weight) weight) (g) 26 8.19 12.19 12.28 0.09 0.44 5.36 0.86 1.05 19.59 1.74 2.12 39.55 2.29 2.80 52.27

27 6.56 12.16 12.22 0.06 0.30 4.57 0.85 1.29 28.24 1.85 2.82 61.69 2.02 3.08 67.54 28 6.31 12.04 12.13 0.08 0.41 6.56 0.85 1.35 20.54 1.91 3.02 46.02 0.07 0.11 1.65 15.79 29 5.96 12.16 12.22 0.06 0.32 5.32 1.31 2.19 41.18 1.74 2.93 55.00 2.77 4.65 87.48 30 5.89 13.18 13.25 0.07 0.35 5.93 1.31 2.23 37.60 1.47 2.50 42.10 2.15 3.66 61.65 MEAN 5.55 1.04 1.62 29.43 1.74 2.68 48.87 1.86 2.86 54.12

SD 0.67 0.22 0.49 8.74 0.15 0.33 8.28 0.93 1.52 28.66

SE 0.21 0.07 0.16 2.76 0.05 0.10 2.62 0.29 0.48 9.06

31 6.68 12.19 12.24 0.06 0.29 4.34 0.75 1.12 25.89 1.91 2.87 66.02 2.14 3.20 73.76

32 5.06 12.23 12.27 0.03 0.17 3.34 0.85 1.68 50.18 1.94 3.83 114.53 2.11 4.17 124.75 33 7.65 11.98 12.06 0.08 0.40 5.19 0.83 1.08 20.89 1.79 2.34 45.14 2.07 2.71 52.26 31.58 34 6.09 12.32 12.38 0.06 0.29 4.77 1.82 2.99 62.78 1.33 2.19 45.82 2.06 3.39 71.01 35 7.02 13.22 13.28 0.05 0.27 3.86 1.19 1.69 43.78 1.21 1.73 44.76 1.49 2.12 54.83 MEAN 4.30 1.09 1.71 40.70 1.64 2.59 63.25 1.97 3.12 75.32

SD 0.65 0.40 0.69 15.49 0.30 0.72 26.87 0.25 0.69 26.14

SE 0.21 0.13 0.22 4.90 0.10 0.23 8.50 0.08 0.22 8.27

36 8.76 12.31 12.38 0.06 0.32 3.70 0.80 0.92 24.80 1.86 2.12 57.37 2.35 2.68 72.57

37 9.17 13.19 13.22 0.04 0.18 1.97 0.76 0.83 41.95 1.37 1.49 75.88 2.33 2.54 128.83 38 4.91 12.23 12.28 0.05 0.24 4.82 0.80 1.62 33.72 1.37 2.78 57.81 2.41 4.92 102.10 63.17 39 6.43 12.24 12.33 0.08 0.42 6.57 3.08 4.79 72.83 2.31 3.59 54.63 4.26 6.62 100.76 40 5.99 12.09 12.19 0.10 0.52 8.61 2.76 4.60 53.40 2.88 4.80 55.76 3.96 6.62 76.80 MEAN 5.13 1.64 2.55 45.34 1.96 2.96 60.29 3.06 4.67 96.21

SD 2.30 1.05 1.77 16.67 0.58 1.16 7.88 0.86 1.80 20.27

SE 0.73 0.33 0.56 5.27 0.18 0.37 2.49 0.27 0.57 6.41

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Table B: Raw bio-accumulation data for 4,4’-DDE, 4,4’-DDD and 4,4’-DDT extracted from the water of the exposure tanks containing exposure concentrations from 1.97 ng/mL to 7.90 ng/mL, used during the acute exposure experiments.

Conc. Conc. Time (hrs) 4,4'-DDE 4,4'-DDD 4,4'-DDT Time (hrs) 4,4'-DDE 4,4'-DDD 4,4'-DDT (ng/mL) (ng/mL) 0 0.34 4.68 0.34 0 0.34 4.99 3.46 24 0.34 4.63 3.46 24 0.34 4.08 0.34 48 0.34 3.81 0.34 48 0.34 4.33 0.34 72 0.34 3.13 0.34 72 1.24 4.16 0.34 96 0.34 3.06 0.34 96 0.34 2.95 0.34

Tank 3.1: 1.97 MEAN 0.34 3.86 0.97 Tank 3.2: 1.97 MEAN 0.52 4.10 0.97 SD 0.00 0.70 1.25 SD 0.36 0.66 1.25 SE 0.00 0.31 0.56 SE 0.16 0.29 0.56 0 9.30 9.87 9.29 0 8.40 9.58 8.90 24 7.42 8.46 8.15 24 7.71 8.52 8.88 48 3.83 8.18 8.78 48 6.09 6.89 8.29 72 5.31 6.24 7.34 72 4.23 5.37 6.58 96 3.33 4.47 5.50 96 3.52 5.16 5.70

Tank 4.1: 3.95 MEAN 5.84 7.44 7.81 Tank 4.2: 3.95 MEAN 5.99 7.10 7.67 SD 2.24 1.89 1.33 SD 1.90 1.73 1.30 SE 1.00 0.84 0.59 SE 0.85 0.77 0.58 0 3.32 4.84 5.95 0 3.26 4.90 6.06 24 2.93 4.15 5.51 24 2.77 3.95 5.33 48 2.27 3.96 4.67 48 2.45 3.73 4.90 72 2.30 3.66 4.83 72 1.71 3.56 4.76 96 2.05 3.66 4.64 96 2.02 3.51 4.60

Tank 5.1: 7.90 MEAN 2.57 4.06 5.12 Tank 5.2: 7.90 MEAN 2.44 3.93 5.13 SD 0.47 0.44 0.52 SD 0.55 0.51 0.52 SE 0.21 0.19 0.23 SE 0.24 0.23 0.23

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Table B continued: Table B: Raw bio-accumulation data for 4,4’-DDE, 4,4’-DDD and 4,4’-DDT extracted from the water of the exposure tanks containing exposure concentrations from 15.79 ng/mL to 63.17 ng/mL, used during the acute exposure experiments.

Conc. Conc. Time (hrs) 4,4'-DDE 4,4'-DDD 4,4'-DDT Time (hrs) 4,4'-DDE 4,4'-DDD 4,4'-DDT (ng/mL) (ng/mL) 0 2.05 3.42 6.20 0 2.18 3.64 7.43 24 2.26 3.62 5.26 24 2.22 3.55 5.37

48 2.12 3.59 5.20 48 1.71 3.39 4.79 72 2.04 6.81 4.81 72 2.00 3.43 4.84

96 1.92 3.26 4.67 6.2: 15.79 96 1.92 3.28 4.73 MEAN Tank 6.1: 15.79 MEAN 2.08 4.14 5.23 Tank 2.01 3.46 5.43 SD 0.11 1.34 0.53 SD 0.18 0.12 1.02 SE 0.05 0.60 0.24 SE 0.08 0.06 0.46 0 25.39 23.42 10.77 0 19.77 18.64 14.50 24 13.15 13.42 10.75 24 12.58 12.53 11.29

48 5.70 8.62 10.06 48 9.36 10.00 9.75 72 4.81 6.69 6.94 72 4.12 5.83 8.35 96 5.18 6.43 6.76 96 4.95 6.32 6.73 MEAN MEAN Tank 7.1: 31.58 10.85 11.71 9.06 Tank 7.2: 31.58 10.16 10.66 10.12 SD 7.90 6.37 1.82 SD 5.70 4.69 2.66 SE 3.53 2.85 0.81 SE 2.55 2.10 1.19 0 4.31 5.55 8.21 0 4.03 5.07 16.34 24 3.43 4.51 8.15 24 3.21 4.04 8.73

48 2.44 3.89 5.63 48 2.56 4.02 5.95 72 2.36 3.82 5.09 72 2.16 3.47 5.14 96 1.99 3.38 4.84 96 1.97 3.41 5.69 MEAN MEAN Tank 8.1: 63.17 2.91 4.23 6.39 Tank 8.2: 63.17 2.79 4.00 8.37 SD 0.85 0.75 1.49 SD 0.75 0.60 4.17 SE 0.38 0.34 0.67 SE 0.34 0.27 1.87

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Table C: In situ water quality parameters read at 24 hour intervals during the acute exposure experiment. The readings from the control tank to the 1.97 ng/mL concentration tank are listed in this section of the table.

Conc. D.O O2 EC Temp. TDS Conc. D.O O2 EC Temp. TDS pH pH (ng/mL) (mg/L) (%) (µS/cm) (°C) (mg/L) (ng/mL) (mg/L) (%) (µS/cm) (°C) (mg/L) 7.71 3.81 48.2 262 26.1 133 7.71 3.81 48.2 262 26.1 133

7.63 5 61.6 259 25.8 131 7.63 5 61.6 259 25.8 131 7.19 7.69 93.3 259 25.4 134 7.8 7.71 94.6 268 25.6 134

Control 7.93 6.66 81.6 262 25.4 132 Control 7.68 6.41 78.9 266 25.8 134 6.92 7.7 94.5 268 28.5 135 6.94 7.51 92.6 266 25.9 135 MEAN 7.48 6.17 75.84 262.00 26.24 133.00 MEAN 7.55 6.09 75.18 264.20 25.84 133.40 SD 0.41 1.72 20.33 3.67 1.30 1.58 SD 0.35 1.67 20.04 3.63 0.18 1.52 SE 0.18 0.77 9.09 1.64 0.58 0.71 SE 0.16 0.75 8.96 1.62 0.08 0.68

7.98 3.76 46.6 754 26.1 377 7.98 3.76 46.6 754 26.1 377 7.84 4.89 60.2 752 25.8 377 7.84 4.89 60.2 752 25.8 377

7.48 7.57 93.4 746 25.8 373 Control 7.62 8.86 108.2 882 25.3 442 7.97 6.31 77.7 740 25.8 370 7.66 6.34 77.5 873 25.4 435 Solvent ControlSolvent 6.93 7.2 88.9 731 25.9 376 Solvent 6.61 6.95 85.7 868 25.8 436 MEAN 7.64 5.95 73.36 744.60 25.88 374.60 MEAN 7.54 6.16 75.64 825.80 25.68 413.40 SD 0.45 1.60 19.69 9.37 0.13 3.05 SD 0.54 1.96 23.70 66.65 0.33 33.34 SE 0.20 0.72 8.81 4.19 0.06 1.36 SE 0.24 0.88 10.60 29.81 0.15 14.91 7.82 4.27 53.1 660 26.2 328 7.82 4.27 53.1 660 26.2 328

7.81 4.81 59.4 675 26 329 7.81 4.81 59.4 675 26 329 7.91 6.64 82.3 659 25.8 330 7.7 6.44 79.1 752 25.4 377 Tank 3.1: Tank 3.2:

1.97 ng/mL 7.85 6.55 80.9 662 26.1 332 1.97 ng/mL 7.74 6.17 75.5 757 25.5 379 6.84 6.52 81 665 26.3 333 6.63 6.43 79.3 755 25.8 379 MEAN 7.65 5.76 71.34 664.20 26.08 330.40 MEAN 7.54 5.62 69.28 719.80 25.78 358.40 SD 0.45 1.13 13.97 6.46 0.19 2.07 SD 0.51 1.01 12.20 48.07 0.33 27.31 SE 0.20 0.50 6.25 2.89 0.09 0.93 SE 0.23 0.45 5.45 21.50 0.15 12.21

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Table C continued: In situ water quality parameters read at 24 hour intervals during the acute exposure experiment. The readings from the 3.95 ng/mL concentration tank to the 15.79 ng/mL are listed in this section of the table.

Conc. D.O O2 EC Temp. TDS Conc. D.O O2 EC Temp. TDS pH pH (ng/mL) (mg/L) (%) (µS/cm) (°C) (mg/L) (ng/mL) (mg/L) (%) (µS/cm) (°C) (mg/L) 7.64 3.81 47.4 992 26.4 497 7.64 3.81 47.4 992 26.4 497

7.61 4.85 59.9 996 26.1 498 7.61 4.85 59.9 996 26.1 498 7.84 6.69 82.6 998 26 499 7.55 7.3 89.8 877 25.8 439 Tank 4.1: Tank 4.2:

3.95 ng/mL 7.51 5.92 73.8 1003 26.5 503 3.95 ng/mL 7.81 5.96 73.4 886 25.8 444 6.68 6.52 81 1003 26.4 503 6.60 6.31 78.1 884 26.2 444 MEAN 7.46 5.56 68.94 998.40 26.28 500.00 MEAN 7.44 5.65 69.72 927.00 26.06 464.40 SD 0.45 1.21 15.01 4.72 0.22 2.83 SD 0.48 1.35 16.44 61.27 0.26 30.29 SE 0.20 0.54 6.71 2.11 0.10 1.26 SE 0.21 0.60 7.35 27.40 0.12 13.54 8.03 5.06 61.8 1125 25.3 560 8.03 5.06 61.8 1125 25.3 560

7.97 7.93 96.9 1126 25.4 616 7.86 8.63 105.2 1235 25.3 616 5.2: 7.12 6.59 80.5 1128 25.4 560 7.98 6.39 78.1 1125 25.4 612 Tank 5.1: Tank

7.90 ng/mL 7.4 6.97 85.3 1131 25.6 563 7.90 ng/mL 7.17 7.55 92.7 1234 25.8 615 7.22 7.74 94.7 1134 25.3 568 7.65 7.16 87.9 1239 25.3 618 MEAN 7.55 6.86 83.84 1128.80 25.40 573.40 MEAN 7.74 6.96 85.14 1191.60 25.42 604.20 SD 0.43 1.15 14.03 3.70 0.12 24.04 SD 0.35 1.33 16.29 60.83 0.22 24.80 SE 0.19 0.51 6.28 1.66 0.05 10.75 SE 0.16 0.60 7.29 27.20 0.10 11.09 7.95 4.84 59.5 1117 25.7 557 7.95 4.84 59.5 1117 25.7 557

7.88 7.55 92.6 1121 25.7 559 7.85 8.56 104 1494 25.2 745 7.86 6.33 78 1113 25.9 556 7.81 6.31 76.9 1494 25.3 747

Tank 6.1: 7.36 7.17 88.4 1123 25.9 560 Tank 6.2: 7.28 7.01 85.7 1488 25.4 743 15.79 ng/mL 15.79 ng/mL 7.91 7.25 89.4 1128 26.1 565 8 7.15 88.1 1499 25.8 748 MEAN 7.79 6.63 81.58 1120.40 25.86 559.40 MEAN 7.78 6.77 82.84 1418.40 25.48 708.00 SD 0.24 1.10 13.50 5.73 0.17 3.51 SD 0.29 1.35 16.31 168.53 0.26 84.43 SE 0.11 0.49 6.04 2.56 0.07 1.57 SE 0.13 0.61 7.29 75.37 0.12 37.76

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Table C continued: In situ water quality parameters read at 24 hour intervals during the acute exposure experiment. The readings from the 31.58 ng/mL concentration tank to the 63.17 ng/mL are listed in this section of the table.

Conc. D.O O2 EC Temp. TDS Conc. D.O O2 EC Temp. TDS pH pH (ng/mL) (mg/L) (%) (µS/cm) (°C) (mg/L) (ng/mL) (mg/L) (%) (µS/cm) (°C) (mg/L) 7.95 4.8 59.2 1279 25.8 638 7.95 4.8 59.2 1279 25.8 638

7.56 5.7 70.4 1283 25.6 640 7.79 6.88 83.3 1359 25.2 678 7.78 6.53 80.9 1277 26.1 637 7.6 6.14 74.8 1349 25.3 672

Tank 7.1: 7.03 6.52 80.8 1267 26.2 631 Tank 7.2: 7.07 6.47 79.3 1332 25.5 665 31.58 ng/mL 31.58 ng/mL 7.66 6.73 83.7 1254 25.3 629 7.68 6.64 82.3 1329 26.3 667 MEAN 7.60 6.06 75.00 1272.00 25.80 635.00 MEAN 7.62 6.19 75.78 1329.60 25.62 664.00 SD 0.35 0.81 10.19 11.66 0.37 4.74 SD 0.33 0.82 9.84 30.85 0.44 15.38 SE 0.16 0.36 4.55 5.22 0.16 2.12 SE 0.15 0.37 4.40 13.80 0.20 6.88 7.85 4.72 58.4 1169 26.1 584 7.85 4.72 58.4 1169 26.1 584

7.87 7.39 90.6 1186 26 591 7.77 7.26 89.2 1101 25.6 549 7.78 5.99 74.4 1181 26.3 589 7.76 5.97 73.5 1092 25.8 544

Tank 8.1: 7.24 6.4 79.2 1172 26.3 585 Tank 8.2: 6.31 6.39 78.7 1078 25.9 537 63.17 ng/mL 63.17 ng/mL 8.02 6.66 82.6 1175 25.9 588 7.59 6.35 78.6 1075 26.1 536 MEAN 7.75 6.23 77.04 1176.60 26.12 587.40 MEAN 7.46 6.14 75.68 1103.00 25.90 550.00 SD 0.30 0.99 11.98 6.88 0.18 2.88 SD 0.65 0.92 11.22 38.37 0.21 19.74 SE 0.13 0.44 5.36 3.08 0.08 1.29 SE 0.29 0.41 5.02 17.16 0.09 8.83

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Table D: Raw biomarker data of all biomarkers analysed in the liver and muscle tissue of Synodontis zambezensis following an acute exposure to 4,4’-DDT. This section of the table included the concentrations from the control to exposure concentration 3.95 ng/mL. Liver Tissue Gill Tissue

Conc. AChE SOD CYP450 MDA PC CAT GSH SOD PC CAT GSH (ng/mL) 0.006 0.809 0.0005 0.267 379.789 740.932 1.877 1.721 150.384 51.8734 1.303

0.007 0.429 0.0004 0.188 552.308 ND 1.892 1.525 135.139 53.7125 1.245 0.004 0.711 0.0004 0.325 416.239 99.552 2.080 1.779 182.149 57.7615 1.724 Control 0.006 0.777 0.0004 0.434 215.168 ND 1.889 1.921 139.797 52.8368 1.892 0.007 1.231 0.0005 0.418 359.638 ND 1.479 2.064 273.248 71.402 1.569 MEAN 0.006 0.792 0.0004 0.326 384.629 420.242 1.843 1.802 176.143 57.5172 1.547 SD 0.001 0.257 0.0000 0.092 108.148 320.690 0.197 0.183 51.2473 7.22607 0.246 SE 0.000 0.091 0.0000 0.033 38.236 143.417 0.070 0.058 16.2058 2.28508 0.078 0.00584 0.91510 0.00062 0.34023 270.72358 238.74332 2.41131 2.54827 226.01908 39.23278 2.05897 0.00467 1.14161 0.00045 0.16604 208.57433 77.37572 2.20574 2.20478 157.46835 42.27417 1.91301

0.00444 1.02854 0.00051 0.37017 218.21934 146.05752 1.53674 2.48255 427.71429 115.07785 2.02858 1.97 0.00442 0.94552 0.00042 0.15804 124.27680 161.43421 2.35782 3.72213 186.24506 42.59164 1.84122 0.00705 0.86461 0.00050 0.09452 247.15010 134.10707 2.78783 2.24365 279.27301 50.12843 1.92993 MEAN 0.00528 0.97907 0.00050 0.22580 213.78883 151.54357 2.25989 2.64027 255.34396 57.86098 1.95434 SD 0.00115 0.10861 0.00008 0.12179 55.70854 58.17349 0.45747 0.55690 95.39221 28.83287 0.07945 SE 0.00036 0.03435 0.00002 0.03851 17.61659 18.39607 0.14466 0.17611 30.16567 9.11776 0.02512 0.00401 0.86482 0.00046 0.10357 273.56651 ND 1.93230 2.20941 138.38298 41.48659 1.20381 0.00586 0.19930 0.00059 0.10965 443.70066 360.25450 2.25059 2.00578 165.96295 53.56712 1.58866

0.00461 0.83262 0.00057 0.15832 206.48363 183.17501 1.79353 1.80862 216.55874 52.92607 1.49069 3.95 0.00589 0.72266 0.00054 0.12708 229.68601 ND 2.49485 2.55264 125.23517 40.59116 1.83696 0.00516 2.52102 0.00059 0.06307 196.07988 158.89380 2.35950 1.68875 279.19132 62.98526 2.27870 MEAN 0.00511 1.02808 0.00055 0.11234 269.90334 234.10777 2.16616 2.05304 185.06623 50.31124 1.67976 SD 0.00081 0.87663 0.00006 0.03480 101.62225 109.91880 0.29415 0.30618 56.52491 8.37166 0.36167 SE 0.00026 0.27721 0.00002 0.01100 32.13578 34.75938 0.09302 0.09682 17.87475 2.64735 0.11437

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Table D continued: Raw biomarker data of all biomarkers analysed in the liver and muscle tissue of Synodontis zambezensis following an acute exposure to 4,4’-DDT. This section of the table included the concentrations exposure concentration 7.90 ng/mL to exposure concentration 15.79 ng/mL. Liver Tissue Gill Tissue

Conc. AChE SOD CYP450 MDA PC CAT GSH SOD PC CAT GSH (ng/mL) 0.00624 1.17751 0.00045 0.24981 343.07320 87.25889 2.00895 6.30622 232.35505 18.26265 1.46652 0.00568 0.93629 0.00051 0.21061 324.11143 290.13826 2.09752 3.93541 261.22734 44.33057 1.67921

0.00472 0.83305 0.00049 0.33697 251.64349 182.33067 2.50675 2.53778 170.02457 22.06199 2.12433 7.90 0.00614 0.72477 0.00053 0.68612 218.67354 152.72318 1.21847 2.48004 273.60533 19.96475 1.20544 0.00598 0.97997 0.00052 1.82232 226.71342 40.66175 1.24212 2.08741 152.50432 31.43307 2.06097 MEAN 0.00575 0.93032 0.00050 0.66117 272.84302 150.62255 1.81476 3.46937 217.94332 27.21061 1.70730 SD 0.00061 0.16979 0.00003 0.67569 57.16735 95.63052 0.56568 1.55057 48.49316 9.69535 0.34918 SE 0.00019 0.05369 0.00001 0.21367 18.07790 30.24103 0.17888 0.49033 15.33488 3.06594 0.11042 0.00559 1.04037 0.00042 0.64605 227.86639 ND 1.02630 1.77580 124.27046 23.74498 1.77778

0.00786 0.77003 0.00050 0.21946 201.09656 163.78457 1.37012 1.53863 146.71435 48.35609 2.01503 0.00962 0.91223 0.00043 0.27459 295.86631 278.89465 1.61884 2.15338 373.01980 40.17020 2.08422 15.79 0.00559 1.43402 0.00049 0.71853 300.65920 122.07693 1.36460 2.49322 296.51936 65.07290 1.76999 0.00399 1.02792 0.00045 0.11404 183.02583 204.19279 1.57831 2.57918 134.50820 45.48614 2.24122 MEAN 0.00653 1.03691 0.00046 0.39453 241.70286 192.23724 1.39163 2.10804 215.00643 44.56606 1.97765 SD 0.00221 0.24726 0.00004 0.27016 54.06672 66.79438 0.23512 0.40175 100.98453 13.33103 0.18182 SE 0.00070 0.07819 0.00001 0.08543 17.09740 21.12224 0.07435 0.12704 31.93411 4.21564 0.05750

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Table D continued: Raw biomarker data of all biomarkers analysed in the liver and muscle tissue of Synodontis zambezensis following an acute exposure to 4,4’-DDT. This section of the table included the concentrations exposure concentration 31.58 ng/mL to exposure concentration 63.17 ng/mL. Liver Tissue Gill Tissue

Conc. AChE SOD CYP450 MDA PC CAT GSH SOD PC CAT GSH (ng/mL) 0.00414 1.34691 0.00050 0.22584 404.81144 141.12520 1.52479 3.25280 259.06492 26.63318 2.14745

0.00277 2.15329 0.00058 0.27801 245.69512 209.70264 1.08524 3.24286 217.79629 29.46645 1.91300 0.00325 0.84696 0.00062 0.23167 244.60641 143.33726 1.14066 4.30873 310.48790 26.37135 1.94688 31.58 0.00341 1.14807 0.00040 0.32911 231.70890 601.89529 1.11496 5.36631 234.09346 28.25107 2.11698 0.00526 1.23945 0.00048 0.30758 220.28436 ND 1.42965 4.14782 223.42342 30.72256 2.30777 MEAN 0.00377 1.34693 0.00052 0.27444 269.42125 274.01510 1.25906 4.06370 248.97320 28.28892 2.08642 SD 0.00097 0.48767 0.00009 0.04553 76.39677 220.89057 0.20292 0.78680 33.85882 1.65705 0.14367 SE 0.00031 0.15422 0.00003 0.01440 24.15878 69.85173 0.06417 0.24881 10.70710 0.52401 0.04543 0.00517 1.30873 0.00044 0.45675 270.62773 342.25756 1.38717 2.63866 232.87982 21.57882 2.38728

63.17 0.00544 1.83836 0.00042 0.29450 216.54602 142.94056 1.28187 3.43929 228.24093 30.75209 6.94041 MEAN 0.00531 1.57354 0.00043 0.37563 243.58687 242.59906 1.33452 3.03898 230.56037 26.16545 4.66384 SD 0.00019 0.37450 0.00001 0.11473 38.24154 140.93840 0.07446 0.40032 2.31945 4.58664 2.27657 SE 0.00009 0.16748 0.00001 0.05131 17.10214 63.02957 0.03330 0.12659 0.73347 1.45042 0.71991

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Table E: In situ water quality data recorded every 24 hours during the acute juvenile toxicity exposure experiments. The control to exposure concentration 2.50 ng/mL readings are listed in this section of the table. Conc. EC (µs/cm) Temp (ºC) pH O2 (%) (ng/mL) 1103 23.7 8.2 48.2

1121 23.6 7.9 61.6 1127 23.5 7.8 93.3

Control 1136 23.6 8 81.6 1147 23.4 8.1 94.5 MEAN 1126.8 23.56 8 75.84 SD 14.78377489 0.10198039 0.141421356 18.18225509 SE 6.611505124 0.045607017 0.063245553 8.131351671

1108 23.7 8.3 46.6 1118 23.5 8 60.2 1123 23.9 7.9 93.4 1130 23.8 7.9 77.7

Solvent ControlSolvent 1143 23.5 8 88.9 MEAN 1124.4 23.68 8.02 73.36 SD 11.7405281 0.16 0.146969385 17.61369921 SE 5.250523783 0.071554175 0.065726707 7.877085756 1116 23.7 8.1 53.1 1117 23.7 7.8 59.4

1126 23.8 7.7 82.3 0.60 1131 23.6 7.7 80.9 1138 23.4 7.8 81 MEAN 1125.6 23.64 7.82 71.34 SD 8.357032966 0.1356466 0.146969385 12.49073256 SE 3.737378761 0.060663004 0.065726707 5.586025421 1107 23.7 8.3 47.4 1115 23.6 7.9 59.9

1121 23.8 7.8 82.6 1.20 1126 23.6 7.8 73.8 1131 23.6 7.9 81 MEAN 1120 23.66 7.94 68.94 SD 8.390470785 0.08 0.18547237 13.4264813 SE 3.752332608 0.035777088 0.082945765 6.004504975 1116 24 8.1 48.2 1117 23.7 7.7 61.6

1122 23.8 7.7 94.6 2.50 1128 23.7 7.7 78.9 1134 23.6 7.8 92.6 MEAN 1123.4 23.76 7.8 75.18 SD 6.8 0.1356466 0.154919334 17.92633816 SE 3.041052449 0.060663004 0.069282032 8.016902145

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Table E continued: In situ water quality data recorded every 24 hours during the acute juvenile toxicity exposure experiments. Exposure concentration 3.70 ng/mL to exposure concentration 2.50 ng/mL readings are listed in this section of the table. Conc. EC (µs/cm) Temp (ºC) pH O2 (%) (ng/mL) 1111 24 8.3 46.6 1116 23.8 7.9 60.2

1121 23.8 7.7 108.2 3.70 1127 23.6 7.8 77.5 1133 23.7 7.8 85.7 MEAN 1121.6 23.78 7.9 75.64 SD 7.787168934 0.132664992 0.20976177 21.19684882 SE 3.482527818 0.059329588 0.093808315 9.479518975 1106 24 8.1 53.1 1118 23.8 7.8 59.4

1124 23.8 7.6 79.1 5.60 1131 23.7 7.7 75.5 1135 23.8 7.8 79.3 MEAN 1122.8 23.82 7.8 69.28 SD 10.22545842 0.09797959 0.167332005 10.90805207 SE 4.572964028 0.043817805 0.074833148 4.878229187 1111 24 8.4 47.4 1117 23.9 7.9 59.9

1121 23.9 7.8 89.8 7.40 1127 23.7 7.8 73.4 1132 23.5 7.9 78.1 MEAN 1121.6 23.8 7.96 69.72 SD 7.364781056 0.178885438 0.224499443 14.70773946 SE 3.293630216 0.08 0.100399203 6.577501045 1113 24 8.1 59.2 1117 23.9 7.7 83.3

1125 23.9 7.5 74.8 8.60 1133 23.8 7.7 79.3 1138 23.8 7.9 82.3 MEAN 1125.2 23.88 7.78 75.78 SD 9.389355675 0.074833148 0.203960781 8.801227187 SE 4.199047511 0.033466401 0.091214034 3.936028455 1109 24 8.3 58.4

1113 24 7.9 89.2 1124 23.8 7.7 73.5 10.50 1130 23.8 7.8 78.7 1136 23.9 7.9 78.6 MEAN 1122.4 23.9 7.92 75.68 SD 10.13114011 0.089442719 0.203960781 10.03780853 SE 4.530783597 0.04 0.091214034 4.489044442

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Table E continued: In situ water quality data recorded every 24 hours during the acute juvenile toxicity exposure experiments. Exposure concentration 12.30 ng/mL readings are listed in this section of the table. Conc. EC (µs/cm) Temp (ºC) pH O2 (%) (ng/mL) 1113 23.9 8 59.2

1116 23.9 7.6 83.3 1125 24 7.5 74.8 12.30 1133 23.8 7.7 79.3 1147 23.9 7.8 82.3 MEAN 1126.8 23.9 7.72 75.78 SD 12.3028452 0.063245553 0.172046505 8.801227187 SE 5.501999636 0.028284271 0.076941536 3.936028455

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Table F: Raw LC50 data including log10 of the exposure concentrations and percentage mortality in each of the exposure concentrations for the three juvenile fish species Synodontis zambezensis, Poecilia reticulata and Oreochromis mossambicus. Conc. Total Exposed Total Dead Mortality Log10 Conc. p=r/n (ng/mL) (n) (r) (%) 0.60 -0.22 10 2 0.2 20

1.20 0.08 10 4 0.4 40 2.50 0.40 10 4 0.4 40 3.70 0.57 10 7 0.7 70 5.60 0.75 10 8 0.8 80 7.40 0.87 10 9 0.9 90 8.60 0.93 10 10 1 100

Synodontis zambezensis 10.50 1.02 10 10 1 100 12.30 1.09 10 10 1 100 0.60 -0.22 10 0 0 0 1.20 0.08 10 0 0 0

2.50 0.40 10 1 0.1 10 3.70 0.57 10 1 0.1 10 5.60 0.75 10 2 0.2 20 7.40 0.87 10 2 0.2 20

Poecilia reticulataPoecilia 8.60 0.93 10 5 0.5 50 10.50 1.02 10 6 0.6 60 12.30 1.09 10 8 0.8 80 0.60 -0.22 10 2 0.2 20

1.20 0.08 10 0 0 0 2.50 0.40 10 1 0.1 10 3.70 0.57 10 3 0.3 30 5.60 0.75 10 7 0.7 70 7.40 0.87 10 9 0.9 90 8.60 0.93 10 9 0.9 90

Oreochromis mossambicus 10.50 1.02 10 10 1 100 12.30 1.09 10 10 1 100

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Appendix B: Published Manuscript

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Figure A: Front page of published article entitled ”Artificial breeding and emryonic development of Synodontis zambezensis (Peters, 1852)” published in Aquaculture Research in July 2018.

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Figure B: Front page of published article entitled ”Baseline bio-accumulation concentrations and resulting oxidative stress in Synodontis zambezensis after an acute laboratory exposure to 4,4’-DDT” published online in Pesticide Biochemistry and Physiology in February 2019.

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Appendix C: Conference Contributions

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CM Volschenk, GP Tweddle, R Greenfield, JHJ van Vuren (2018) Artificial breeding and normal development of Synodontis zambezensis (Peters, 1852), South African Society of Aquatic Scientists 2018, 24-28 June 2018, Cape St Francis Bay Resort, Eastern Cape.

Artificial Breeding and normal development of Synodontis zambezensis (Peters, 1852).

CM Volschenk1, G Tweddle1, R Greenfield1 and JHJ van Vuren1

1Department of Zoology, Kingsway Campus, University of Johannesburg, PO Box 524, Auckland Park, 2006

Synodontis sp. are an important source of food in many African countries including, Nigeria, South Africa, Egypt and Ghana. These fish have a high fat content making them an excellent source of fat in fish-based diets. However, their ability to transfer accumulated lipophilic pollutants to the human population is increased. The species of focus is Synodontis zambezensis and was selected because of its importance to informal communities. This study resulted in the successful breeding of S. zambezensis in captivity following a combination of protocols used in artificial breeding of Clarias gariepinus, Synodontis petricola and Synodontis nigromaculatus. Fish were injected with Aquaspawn® and manual fertilisation of stripped eggs followed. The embryonic development was meroblastic in nature and discoidal cleavage followed as seen in S. nigromaculatus and C. gariepinus. The resulting larvae hatched within 26 hours of fertilisation and the survival rate of fry and subsequent fingerlings was as high as 80%. S. zambezensis has already been successfully used as a bio-indicator species in field studies. This means that successful artificial breeding will allow for their use as an indigenous test organism in laboratory based chemical testing and bio-accumulation of harmful chemicals such as Dichlorodiphenyltrichloroethane (DDT).

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Edwards CM, Ikenaka Y, Saengtienchai A, Ishizuka M, Van Vuren JHJ and Greenfield R (2017) Novel sequencing of genes in a South African sentinel fish species, 9th International Toxicology Symposium in Nigeria, 6-9 September 2017, Constantial Hotel, Benin City, Nigeria.

Novel sequencing of genes in a South African sentinal fish species

Claire M Edwards*1, Yoshinori Ikenaka2,3, Aksorn Saengtienchai2, Mayumi Ishizuka2, Johan HJ van Vuren1 and Richard Greenfield1

1Department of Zoology, Kingsway Campus, University of Johannesburg, PO Box 524, Auckland Park, 2006

2Laboratory of Toxicology, Department of Environmental Veterinary Sciences, Graduate School of Veterinary Medicine, Hokkaido University, N18, W9, Kita-ku, Sapporo 060-0818, Japan

3School of Biological Sciences, Potchefstroom Campus, North West University, X6001, Potchefstroom, 2520

This study aims to sequence genes and housekeeping genes within the species Synodontis zambezensis so that further investigations into the effect of Dichlorodiphenyltrichloroethane (DDT) may be conducted. Genes need to be identified and sequenced within this species since it has been confirmed as an important source of food to rural communities and because effects of pollution at a genetic level have not been assessed. The focus genes include Transferrin, Estrogen Receptor (ER) and the housekeeping gene β-actin. This initial study proved successful in identifying the very important β-actin housekeeping gene, but other genes related to the sexual development of fish such as Choriogenin H and L and Vitellogenin proved very difficult to find. Genes related to the breakdown of toxins within the liver such as Cyp1A1 and Cyp19a, also proved very difficult to find within conserved areas of various Siluriforme fish.

Keywords: Synodontis zambezensis, DDT, housekeeping gene

Introduction Synodontis zambezensis also known as the Brown Squeaker is a catfish belonging to the family Mochokidae, the largest family of Catfish in Africa (Skelton, 2001). The importance of this species was recently discovered in an Assessment of the Lower Phongola River and Floodplain in KwaZulu-Natal (KZN) South Africa (Smit et al., 2016). Since this fish is eaten by the local fishing community (Coetzee et al., 2015),

177 | P a g e the need to understand the effects of pollution on this species has become more important.

Bio-accumulation of Organochlorine pesticides (OCP’s) and resulting biomarker effects were investigated (Smit et al., 2016)), but the effect of pollutants at the gene level has not been assessed. Dichlorodiphenyltrichloroethane (DDT) was the main OCP found in S. zambezensis during the above mentioned investigation and this lipophilic pesticide has been shown to cause changes in the regulation of sex-linked genes like Vitellogenin and Estrogen Receptor (ER) (Larkin et al., 2003). DDT is known to be an Endocrine Disrupting Chemical (EDC) and may cause changes in the sexual development of fish. This may result in decreased spawning and a decline in the overall population of the species (Hutchinson et al., 2006).

In order for successful comparisons on gene expression to be made, it is essential to first identify a “housekeeping” gene within the target species. These genes are usually constitutive genes that are essential for the normal functioning of cells. These genes are present and are expressed under normal conditions, as well as under pathophysiological conditions and for this reason they are used as baselines for the expression of other genes within the organism (Butte et al., 2001). The most common housekeeping genes used in comparative expression are β-actin, g6pd and gapdh (McCurley & Callard, 2008).

The aim of this study was to identify previously unsequenced genes and housekeeping genes within S.zambezensis that can be linked to differing concentrations of DDT and increase our understanding of the endocrine disrupting action of DDT. This report will outline the techniques and protocols used to prepare sampled tissue for genetic analysis as well as some basic findings.

Materials and Methods RNA Extraction/Isolation Total Ribonucleic Acid (RNA) was extracted from the tissues collected from S. zambezensis namely, muscle, liver, kidney and gonads. The tissues were stored in RNAlater® (Sigma-Aldrich co. LLC, South Africa) at -80°C until analysed. The isolation protocol followed was a combination of the TRIzol™Reagent protocol (ThermoFisher Scientific) and the standard protocol supplied in the NucleoSpin® RNA II Kit (Macherey-Nagel GmbH &Co. KG, Düren, Germany). TRIzol™Reagent (700µl) was added to each sample in an Eppendorf tube along with a bead and was

178 | P a g e placed into a shaker for homogenization. Once homogenized, 200µl Chloroform was added and vortexed thoroughly. Samples were then centrifuged at 4°C for 20min at 1200g and the supernatant was pipetted through a NucleoSpin®Blue Ring Filter into a collection Tube. From this stage (step 4) onwards, the standardized NucleoSpin®RNA II Kit protocol was followed. The quality and quantity of extracted RNA was checked spectrophotomically using a Nanodrop ND-1000 (Thermo Scientific, DE, USA) with all A260/280 and A260/230 readings ≥2. Agarose gels were run to visually confirm the presence of required 18s and 28s bands. cDNA Synthesis Extracted RNA was converted to complimentary Deoxyribonucleic Acid (cDNA) using ReverTraAce® (Toyoba, Tokyo, Japan). The protocol was as follows: 2µl of RNA template was denatured at 65°C for 5min and immediately placed on ice. To the denatured RNA, 2µl of 4xDN Master Mix was added, as well as 4µl of Nuclease-Free Water and the mixture was incubated at 37°C for 5min. After incubation the mixture was again placed on ice and 2µl 5xRTMaster Mix II was added. This final mixture with a total volume of 40µl was placed into a BIO-RAD iCycler and the following cycle was run: 37°C for 15min, 50°C for 5min, heat to 98°C for 5min. The quality and quantity of the resulting cDNA was checked with a Nanodrop ND-1000 Spectrophotometer (Thermo Scientific, DE, USA) with all A260/280 and A260/230 readings ≥1.8.

Designing Primers Since there is no current genetic data available for S. zambezensis, the forward and reverse primers used for each of the genes had to be designed through identification and alignment of homologous sequences from other already sequenced fishes. Homologous nucleotide sequences were identified using the National Centre for Biotechnology Inormation (NCBI) and were aligned using ClustalW. Forward and reverse primers were selected along the length of conserved regions.

Real Time PCR Real-time PCR (RT-PCR) for S. zambezensis mRNA levels was performed using the following: Ex Taq, 10xEx Taq Buffer, dNTP (Toboyo, Japan). Once all of the reagents, the forward and reverse primers as well as the cDNA template were added together following the supplier standard concentrations, the mixture was placed into the BIO-RAD iCycler and the following was run: 94°C for 4 mins for a single cycle,

179 | P a g e followed by 94°C for 30 sec, 55°C for 30sec, 72°C for 1 min and 72°C for 4min for 35 cycles. The PCR product was run on an agarose gel to confirm any amplification in the target organism DNA.

On confirmation of amplification, the relevant cDNA products were diluted to the correct concentrations and specifications as described by the Food Assessment and Management Centre (FASMAC, Japan) and sent for sequencing. Once sequences were received back from FASMAC, they were examined using BioEdit Sequence Alignment Editor. The resulting amino acid sequences represent sections of the sequenced target genes.

Results and Discussion Primers initially designed for β-actin proved to be unsuccessful and therefore the forward and reverse primers for Silurius meridionalis (Chinese large-mouth catfish) were applied to the tissue samples (Huange et al., 2010). The primers designed through the alignment of homologous genes for Transferrin and ER proved to be successful. Homologous genes from fish species such as Ictalurus punctatus (Channel catfish), Pelteobagrus fulvidraco (Yellow catfish), Oreochromis niloticus (Nile tilapia), Oncorhynchus mykiss (Rainbow trout) and Oryzias latipes (Japanese rice fish), were used during the alignment stage and primers were selected within the conserved regions of each gene. Sequenced genes from the order Siluriformes were used wherever possible, but since sequenced genes from this order are scarce, fish species from other orders were used as well. All forward and reverse primers used for each of the sequenced genes are listed in Table 1. Table 1: Forward and reverse primer sets compiled for the amplification of β-actin, Transferrin and Estrogen Receptor (ER) genes. Gene Forward (5’-3’) Reverse (5’-3’) β-actin CCATCTCCTGCTCGAAGTCC GCCCATCTACGAGGGTTACG Transferrin GCTACTACGCTGTAGCTGTAG CCGCGTCCTCCAGAGGTTTC ER GATCAGTTAATCATCCTGG CGCACACCTCGCGACGACTTTC

Agarose gels run after RT-PCR showed distinct bands of gene amplification in the PCR products for each primer set of each gene. Figure 1A shows the bands formed from the 100bp standard added to each gel for base pair comparison of amplified genes. The bands that follow show amplification products of the primer pairs for ER

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(1B-4B) and β-actin (1C-4C) and these genes are clearly amplified in all four of the sampled tissues. Figure 2 however, shows clear amplification bands for Transferrin in liver (2A) and kidney (3A) tissue, but in muscle (1A) and gonadal (4A) tissue the bands appear at the incorrect base pair positions, suggesting that a different gene is amplified in this case. Reasons for the differences in position can be linked to the types of tissues amplified and to the fact that Transferrin is a gene related to the binding and transport of iron and so may not be present in muscle and gonadal tissues since these organs are less involved in these physiological functions (Crichton & Charloteaux-Wauters, 1987).

Figure 1: Amplification of PCR products for ER (B) and β-actin (C) in muscle (1), liver (2), kidney (3) and gonadal (4) tissue.

Figure 2: Amplification of PCR products for the Transferrin gene in muscle (1), liver (2), kidney (3) and gonadal (4) tissue.

Successfully amplified genes with satisfactory bands at the correct base pair positions, were sent to FASMAC (Japan) for sequencing. The gene sections sequenced so far are as follows: Β-actin: TCCATCTCCTGCTCGAAGTCCAGAGCAACATAGCACAGCTTCTCCTTGATGTCA CGCACGATCTCACGCTCGGCCGTGGTGGTGAAGCTGTAGCCACGCTCGGTCA GGATCTTCATCAGGTAGTCCGTGAGGTCACGGCCAGCCAGGTCCAGACGCAG GATGGCATGGGGCAGAGCGTAACCCTCGTAGATGGGC

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Transferrin: GCTACTACGCTGTAGCTGTAGTAAAGAAAGGCACTAACTTTGGCTTCAAGGACC TTCGTGGGAAGAAGTCCTGTCACACTGGTTTGGGGAAAACTGCAGGCTGGAAC ATCCCCATTGGCACTCTCCTCTCAAAAAAACAAATTCAATGGGGAGGAATCGAT GAGAAACCTCTGGAGGACGCGG

ER: GATCAGTTAATCATCCTGGAAAGCTCATCTTCTCCCCAGATCTTGTCCTCAGCA GGGATGAGGGCAGCTGTGTGCAGGGACCAGACCCGGTAGAGTTTCTACCCGG TCTCGAAAGTCGTCGCGAGGTGTGCG

It is recommended that primer efficiency testing be completed in order to confirm the efficiency of each primer in the tissues of the focus species and to establish whether more efficient primers need to be designed before qRT-PCR analysis can be completed. This will allow for comparative amplification studies to be completed. Other important sex linked genes will be focussed on in further studies to allow for a more holistic understanding of the endocrine disruption that may be taking place within this species.

Acknowledgments This study was completed with funding from the NRF, grant number UID 92424 Japan/ SA Bi-lateral, GES from the University of Johannesburg and NRF Scarce Skills Development Fund Doctoral Scholarship, grant number SFH150706123266. I would like to thank the Laboratory of Toxicology at Hokkaido University for all of their help and support in completing this work.

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Edwards CM, Van Vuren JHJ and Greenfield R (2017) Preliminary results of Oxidative stress biomarkers in Synodontis zambezensis during an acute exposure to DDT, South African Society of Aquatic Scientists 2017, 25-29 June 2017, Birchwood Hotel and O.R Tambo Conference Centre, Gauteng.

Preliminary results of Oxidative stress biomarkers in Synodontis zambezensis during an acute exposure to DDT.

Edwards CM1, JHJ van Vuren1, Greenfield R1

1Department of Zoology, Kingsway Campus, University of Johannesburg, PO Box 524, Auckland Park, 2006

Oxidative stress within organs of fish has long been an indicator of exposure to Organochlorine Pesticides (OCP’s). The tendency for OCP’s to bio-accumulate within the environment and tissues of organisms’ means that their effects are particularly important to understand. Dichlorodiphenyltrichloroethane (DDT) is an OCP that forms the focus of many studies throughout Africa and South Africa, because of its historic and continued use in the control of the malaria vector, Anopheles sp. Oxidative stress effects are caused by the production of reactive oxygen species (ROS’s) or the deactivation of antioxidant defences within fish. The three most important enzymes for detoxification of ROS’s in all aerobic organisms are Superoxide Dismutase (SOD), Catalase (CAT) and Glutathione (GSH). The inhibition or excitation of these three enzymes are analysed as biomarkers of effect, giving insight into the physiological adaptability of the target species, Synodontis zambezensis. Preliminary results of acute exposures to DDT show no significant differences in oxidative stress reactivity at different exposure concentrations, but trends in the physiological adaptations of S. zambezensis allow for preliminary conclusions to be reached. The closely linked SOD and CAT enzyme-relationship is clearly illustrated, with each group showing changes in reactivity in correlation with the other. Glutathione concentrations are naturally high within the hepatic cells of fish and play an essential role in neutralizing ROS’s. This non-protein compound tends to decrease in concentration during acute exposures and this is clearly illustrated in the preliminary results. Investigations into the chronic effects of DDT under controlled conditions are ongoing.

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