Illicit Drugs - Environmental Occurrence, Fate

Illicit drugs - environmental occurrence, fate

and toxicity

Pandian Govindarasu

B.Sc. (Agriculture), M.Sc. (Agricultural Entomology)

Submitted in fulfilment of the requirements for the degree of

Doctor of Philosophy

Global Centre for Environmental Remediation

Faculty of Science and Information Technology

February 2016

DECLARATION

I declare that:

This thesis presents work carried out by myself and does not incorporate without acknowledgment any material previously submitted for a degree or diploma in any university. To the best of my knowledge it does not contain any materials previously published or written by another person except where due reference is made in the text; and all substantive contributions by others to the work presented, including jointly authored publications, are clearly acknowledged.

Pandian Govindarasu

Signed Date: 19 February 2016

ACKNOWLEDGEMENTS It would not have been possible to complete my thesis without the guidance of my supervisors, help from CERAR staff and friends, and support from my wife and family. First and foremost, I wish to thank Professor Mallavarapu Megharaj my principle supervisor for his guidance, caring, patience and providing me with an excellent atmosphere for doing my research work. I thank him once again for his continuous motivation and belief in me to carry out this novel PhD topic in one of the best environmental science research laboratory. I would like to express my deepest gratitude to my associate supervisor, Professor Ravi Naidu for giving me an opportunity to carry out this research project under his supervision. I seriously appreciate his suggestions, advices and comments all through my research. I also thank Cooperate Research Centre for Contamination Assessment and Remediation of the Environment (CRC CARE) for supporting my research and providing me with scholarship. I take this opportunity to thank to Dr. Raktim Pal (Co-supervisor), who always willing to help and give his best suggestions throughout my PhD programme. Many thanks to Dr. Paul Kirkbride for his advice and support to finish my research work successfully on time am truly thankful to Dr. Paul Pigue for providing me with the test chemicals and technical advice entire of my PhD candidature. Special thanks goes to Dr. Logesh and Dr. Prasath (for their help with proof reading, Daphnia study, earthworm photograph, Minitab etc.,), Dr. Raja, Vilma and Peter (for their help with analytical instruments), Dr. Kannan (for his help with sample collection and comet assay), Dr. Thavamani (for his help with earthworms studies and materials), Srinithi (for her help with comet assay). I would also like to thank Suresh, Renga (Jeffries), Vidhya, Vimal, Anitha, Kavitha, Saranya and staff and students at CERAR and CRC CARE for their timely help to complete my research work in a successful manner. Above all, I would like to thank my wife (Amudha) and my sons (Thaarik and Nirai) for their invaluable love, personal support, sacrifice, and great patience at all times. Last, but by no means least, my heartfelt gratitude to my parents, brother and sister have given me their unequivocal support throughout, as always, for which my mere expression of thanks does not sufficient.

TABLE OF CONTENTS

LIST OF FIGURES i LIST OF TABLES iv ABSTRACT vii CHAPTER ONE 1 INTRODUCTION 1 CHAPTER TWO 6 ILLICIT DRUGS – ENVIRONMENTAL OCCURRENCE, FATE, AND 6 TOXICITY Abstract 6 2.1 Introduction 6 2.2 Environmental occurrence of illicit drugs and their metabolites 8 2.2.1 Wastewater 11 2.2.2 Surface water 12 2.2.3 Sewage sludge and bio- solids 13 2.2.4 Atmospheric air 13 2.3 Environmental fate of illicit drugs 14 2.3.1 Water 14 2.3.2 Soil/sediments/sludge 15 2.4 Toxicity of illicit drugs 16 2.4.1 Toxic effects on aquatic organisms 17 2.4.2 Toxic effects on animals 18 2.4.2.1 Cocaine 18 2.4.2.2 MAP and MDMA 20 2.4.2.2.1 Neurotoxicity 21 2.4.2.2.2 Apoptosis and oxidative stress 21 2.4.2.2.3 Behavioural effects 22 2.4.3 Toxic effects on humans 22 2.4.3.1 Cocaine 22 2.4.3.2 MAP and MDMA 23 2.4.4 Toxic effects on plants 24 2.5 Conclusion 24 2.6 References 31

CHAPTER THREE 48 ILLICIT DRUGS - EMERGING CONTAMINANTS IN AN URBAN 48 ENVIRONMENT Abstract 48 3.1 Introduction 48 3.2 Materials and methods 50 3.2.1 Reagents and materials 50 3.2.2 Sample collection and analysis 50 3.2.3 Analysis of illicit drugs and quality control 50 3.3 Results and discussion 51 3.4 Conclusion 53 3.5 References 57 CHAPTER FOUR 65 SORPTION AND DESORPTION PATTERNS OF AMPHETAMINE-TYPE 65 SUBSTANCES (ATS) in DIFFERENT SOILS – THE INFLUENCE OF SOIL PROPERTIES Abstract 65 4.1 Introduction 65 4.2 Materials and methods 66 4.2.1 Chemicals 66 4.2.2 Soils 66 4.2.3 Sorption and desorption test 67 4.2.4 Compounds extraction and analysis 69 4.2.5 Calculation of sorption parameters 69 4.3 Results and discussion 70 4.3.1 Sorption kinetics 70 4.3.2 Sorption isotherms 71 4.3.3 Sorption parameters 73 4.3.4 Correlation of sorption coefficient (Kd) with major soil properties 76 4.3.5 Desorption 77 4.4 Conclusion 79 4.5 References 80 CHAPTER FIVE 82

DEGRADATION OF COCAINE IN SOILS AND ITS ADVERSE EFFECTS ON 82 EARTHWORMS (EISENIA FETIDA) Abstract 82 5.1 Introduction 82 5.2 Materials and methods 83 5.2.1 Chemicals 83 5.2.2 Organisms 84 5.2.3 Soils 84 5.2.4 Cocaine degradation in soil 84 5.2.5 Toxicity assay 85 5.2.6 Lipids estimation 85 5.2.7 Total antioxidant capacity 85 5.2.8 Lipid peroxidation 86 5.2.9 Comet assay 86 5.2.10 Chemical analysis 87 5.3 Results and discussion 87 5.3.1 Cocaine degradation (non-sterile and sterile) 87 5.3.2 Impact of cocaine on earthworms’ weight and total lipids content 90 5.3.3 Effect of cocaine on earthworms’ antioxidant capacity 91 5.3.4 Effect of cocaine on lipid peroxidation of earthworms 93 5.3.5 Cocaine induced DNA damage 94 5.4 Conclusion 96 5.5 References 97 CHAPTER SIX 101 METHAMPHETAMINE (MAP) TOXICITY TO EARTHWORMS (EISENIA 101 FETIDA) FOLLOWING SOIL EXPOSURE Abstract 101 6.1 Introduction 101 6.2 Materials and methods 102 6.2.1 Chemicals and reagents 102 6.2.2 Test soils 103 6.2.3 Test species 103 6.2.4 Toxicity assay 103 6.2.5 Bioaccumulation test 103 6.2.6 Reproduction test 103

6.2.7 Morphological and behavioural test 103 6.2.8 Comet assay 104 6.2.9 Chemical analysis 104 6.3 Results and discussion 104 6.4 Conclusion 110 6.5 References 112 CHAPTER SEVEN 115 OVERDOSES OF PSEUDOEPHEDRINE (PSE) CHRONIC EXPOSURE TO 115 EARTHWORMS – IMPACTS ON LIFE PARAMETERS AND DNA Abstract 115 7.1 Introduction 115 7.2 Materials and methods 116 7.2.1 Soil collection and preparation 116 7.2.2 Reagents 116 7.2.3 Earthworms 117 7.2.4 Toxicity and bioaccumulation test 117 7.2.5 Reproduction test 117 7.2.6 Morphological and behavioural test 117 7.3 Results and discussion 117 7.4 Conclusion 122 7.5 References 124 CHAPTER EIGHT 127 ACUTE TOXICITY AND GENOTOXICITY OF AMPHETAMINE-TYPE 127 STIMULANT METHAMPHETAMINE AND ITS PRECURSOR PSEUDOEPHEDRINE TO DAPHNIA CARINATA Abstract 127 8.1 Introduction 127 8.2 Materials and methods 129 8.2.1 Test organism and culture condition 129 8.2.2 Test compounds 129 8.2.3 Test water 129 8.2.4 Acute toxicity test 129 8.2.5 Chemical stability and analysis of MAP and PSE 130 8.2.6 Comet assay 130 8.3 Results and discussion 131

8.3.1 Physico-chemical properties of water 131 8.3.2 Acute toxicity of test chemicals 133 8.3.3 Genotoxicity of MAP and PSE to D. carinata 135 8.4 Conclusion 137 8.5 References 139 CHAPTER NINE 143 ACUTE AND GENO-TOXICITY OF COCAINE AND MDMA TO DAPHNIA 143 CARINATA Abstract 143 9.1 Introduction 143 9.2 Materials and methods 145 9.2.1 Organism and culture condition 9.2.2 Test compounds 9.2.3 Test water 9.2.4 Acute toxicity test 9.2.5 Chemical stability of cocaine and MDMA 9.2.6 Comet assay 9.2.7 Chemical analysis 9.3 Results and discussion 145 9.3.1 Physico-chemical properties of water 145 9.3.2 Acute toxicity of test chemicals 146 9.3.3 Stability of cocaine and MDMA 148 9.3.4 Genotoxicity of cocaine and MDMA to D. carinata 149 9.4 Conclusion 152 9.5 References 153 CHAPTER TEN 156 PHYTOTOXICITY OF ILLICIT DRUGS TO LEMNA MINOR L. 156 Abstract 156 10.1 Introduction 156 10.2 Materials and methods 158 10.2.1 Chemicals 158 10.2.2 Cultivation and growth of L. minor L. 158 10.2.3 Growth parameters 158 10.2.4 Chlorophyll and free proline estimation 159 10.2.5 Bioaccumulation of illicit drugs 160

10.2.6 Chemical analysis 160 10.3 Results and discussion 160 10.3.1 Effect of illicit drugs on growth parameters (frond number, RGR and fresh 160 weight) 10.3.2 Effect of illicit drugs on inhibition of growth 165 10.3.3 Illicit drugs tissue accumulation 169 10.3.4 Chlorophyll and proline content 170 10.4 Conclusion 172 10.5 References 173 CHAPTER ELEVEN 176 SUMMARY AND CONCLUSIONS 176 11.1 This research has demonstrated that 177 11.2 Propositions for future research 178

LIST OF FIGURES

Figure 3.1 Geographical locations of the sampling points…………………………. 55 Figure 4.1 Sorption kinetics of MAP & MDMA in 3 experimental soils…………… 70 Figure 4.2 A plot for the sorbed amount (μg g-1) of MAP & MDMA in 3 72 experimental soils as a function of equilibrium concentration (μg mL-1) Figure 4.3 Desorption pattern of MAP & MDMA in 3 experimental soils…………... 77 Figure 5.1 Non-sterile degradation of cocaine in three test soils…………………… 85 Figure 5.2 Sterile degradation of cocaine in three test soils……………………...... 88 Figure 5.3 Effect of cocaine on earthworms’ weight and lipids after 28 days soil 90 exposure. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. Different letters show statistically difference at p < 0.05 (ANOVA)…………………………...... Figure 5.4 Effect of cocaine on total antioxidant capacity of earthworm after 28 92 days soil exposure in soil………………………………...... Figure 5.5 Effect of cocaine on earthworms lipid peroxidation after 28 days soil 94 exposure………………………………...... Figure 5.6 DNA damage induced by cocaine in E. fetida. Results are expressed 95 as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA & olive tail movement - ANOVA, p< 0.05) …………………. Figure 5.7 DNA damage in E. fetida exposed to cocaine (a). Control with no or 96 minimal DNA damage (b). DNA damage in cocaine exposed E. fetida in soil………………………………………………………………………… Figure 6.1 MAP effect on earthworm weight changes (%) over control……………. 108 Figure 6.2 MAP effects on earthworm reproduction (%) over control……………… 108 Figure 6.3 Effects of MAP in earthworm following 28 days exposure in soil. (a) 109 Control. (b) Earthworm coiling (20 mg kg-1). (c) Dehydrated earthworm (50 mg kg-1). (d) Cuticle damage and fragmentation (100 mg kg-1) ……. Figure 6.4 DNA damage induced by MAP in earthworms. Results are expressed 110 as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA - ANOVA, p< 0.05 & olive tail movement - Dunett test, p < 0.05) ………………………………………………………………………. Figure 6.5 DNA damage in MAP exposed earthworms (E. fetida) as analysed by 110 the comet assay. (a). Control with no or minimal DNA migrating into the tail region. (b). MAP 5 mg kg-1exposed worms DNA migrating into

the tail region as a result of strand breakage……………………………………………………………………. Figure 7.1 PSE effects on earthworms weight changes (%) over control……….... 118 Figure 7.2 PSE effects on earthworms reproduction (%) over control……………. 121 Figure 7.3 DNA damage induced by PSE in earthworms. Results are expressed 122 as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA - ANOVA, p < 0.05 & olive tail movement - Dunett test, p < 0.05) ………………………………………………………………………. Figure 7.4 DNA damage in PSE exposed earthworms (E. fetida) as analysed by 122 the comet assay. (a). Control with no or minimal DNA migrating into the tail region. (b). PSE exposed (5 mg kg-1) worms DNA migrating into the tail region as a result of strand breakage………………………. Figure 8.1 DNA damage induced by MAP and PSE to D. carinata. Results are 136 expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA & olive tail movement - ANOVA, p < 0.05) …... Figure 8.2 DNA damage in D. carinata exposed to MAP and PSE. (a). Control 137 with no or minimal DNA damage (b) DNA damage in MAP 1 mg L-1 exposed D. carinata and (c) DNA damage in PSE 1 mg L-1 …………… Figure 9.1 DNA damage induced by cocaine and MDMA to D. carinata. Results 151 are expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA & olive tail movement - ANOVA, p< 0.05) …… Figure 9.2 DNA damage in D. carinata exposed to cocaine and MDMA. (a). 152 Control with no or minimal DNA damage (b). DNA damage in cocaine (1 mg L-1) exposed D. carinata (c) DNA damage in 1 mg L-1 of MDMA exposed D. carinata in water……………………………………………… Figure 10.1 L. minor L. growth (total frond number) after exposure to cocaine, 161 MAP, MDMA and PSE for 7d. Bars denote standard deviation (n = 3). Different letters indicate a significant difference (1-way ANOVA, p < 0.05) …………...... Figure 10.2 L. minor L. growth (fresh weight) after exposure to cocaine, MAP, 162 MDMA and PSE for 7d. Bars denote standard deviation (n = 3). Different letters indicate a significant difference (1-way ANOVA, p < 0.05) …………………...... Figure 10.3 L. minor L. growth (relative growth rate) after exposure to A) Cocaine 163 B) MAP C) MDMA and D) PSE for 7d. Bars denote standard deviation

(n = 3). Different letters indicate a significant difference (1-way ANOVA p < 0.001) …………………………………………………………………... Figure 10.4 L. minor L. growth (inhibition of growth based upon fresh weight and 166 frond number) after exposure to A) Cocaine, B) MAP, C) MDMA, D) PSE for 7d…………………………………………………………………...

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LIST OF TABLES

Table 2.1 Selected physico-chemical properties of major illicit drugs and 9 metabolites……………………………………………………………………….

Table 2.2 Illicit drugs’ concentrations in environmental systems………………………. 26 Table 2.3 Aquatic toxicity data for illicit drugs …………………………………………... 27

Table 3.1 Concentration of illicit drugs and their metabolites in wastewaters, sewage 54 sludge and surface waters in Adelaide surrounding areas………………….

Table 3.2 (supplementary) : Sample collection details………………………………….. 61 Table 4.1 Basic physico-chemical properties of the experimental soils………………. 68 Table 4.2 Summary of the sorption parameters of MAP & MDMA in 3 experimental 75 soils………………………………………………………………………………. Table 4.3 Summary of the sorption parameters of MAP & MDMA in 3 experimental 74

soils (Kd, Koc & Gibb’s energy) ………………………………………………...

Table 4.4 Correlation of sorption coefficient (Kd) of MAP & MDMA with soil properties. 76 Table 4.5 Amount of chemical adsorbed and desorbed by soil from MAP (53.6 µg 79 mL-1) and MDMA 54.7 µg mL-1) test solution ………………………………….

Table 5.1 Regression equation, rate constant (k), and half-life (t1/2) values for the 89 degradation of cocaine under non-sterile and sterile conditions…………… Table 6.1 MAP soil concentration, bioaccumulation and biological parameters of 107 earthworms………………………………………………………………………. Table 7.1 PSE soil concentration, bioaccumulation and biological parameters of 120 earthworms………………………………………………………………………. Table 8.1 Physico-chemical properties of waters………………………………………... 132 Table 8.2 Acute toxicity of methamphetamine and pseudoephedrine to Daphnia 134 carinata…………………………………………………………………………...

-1 Table 9.1 LC50 values (mg L ) of cocaine and MDMA to D. carinata tested in water 146 samples…………………………………………………………………………. Table 9.2 Stability of cocaine and MDMA in water samples……………………………. 149 -1 Table 10.1 EC50 values (mg L ) of illicit drugs to L. minor L. on the basis of growth 168 measured as fresh weight and frond number (7d) …………………………… Table 10.2 Illicit drugs concentration in medium and in L. minor L. tissues, exposed in 169 10 mg L-1 for 7d…………………………………………………………………..

Table 10.3 Chlorophyll contents after 7d exposure of L. minor L. to illicit drugs………... 171 Table 10.4 Effect of illicit drugs on Proline content (µM g-1 fresh weight) in L. minor L. 172 7d exposure………………………………………………………………………

LIST OF ABBREVIATIONS ATSs Amphetamine-type stimulants MAP Methamphetamine MAS Methylamphetamine sulphate PSE Pseudoephedrine P2P Phenyl-2-propanone MDMA 3, 4-methylenedioxymethamphetamine COC Cocaine BE Benzoylecgonine EME Ecgonine methyl ester THC Δ9-tetrahydrocannabinol LSD Lysergic acid diethylamide ECs Emerging contaminants

Kd Sorption coefficient DOC Dissolved organic carbon OC Organic carbon RGR Relative growth rate ROS Reactive oxygen species WWTPs Wastewater treatment plants HPLC-MS High performance liquid chromatography - mass spectroscopy

ABSTRACT Illicit drugs are those compounds whose non-medical use is prohibited by international legislation and mainly belong to the classes of opiates, cocaine, cannabis, amphetamines and ecstasy-group substances (UNODC, 2007; Hall et al., 2008). These chemicals have been proven to be toxic to humans and animals in that they cause numerous and potentially precarious side effects. Reports on illicit drugs and their metabolites ending up in the environment have been increasing worldwide. These problems occur mainly due to human consumption and also disposal into sewage networks by the illegal manufacturesrs. Although, the reported environmental concentrations of illicit drugs are low, their potent pharmacological properties and the mixing of illicit drugs along with similar compounds in soil and water could be toxic to non-target organisms and pose a risk to human health. The data available on the environmental occurrence and distribution pattern of illicit drugs in Australia (i.e. South Australia), fate in soil and water and ecotoxicity on aquatic and terrestrial biota are limited given that previously published research has focused mainly only on sewage epidemiology and human health. This thesis focuses on the: (a) occurrence and distribution pattern of illicit drugs and their metabolites in South Australian wastewater (influent and effluent), surface waters, sewage sludge and sediments; (b) fate of illicit drugs in soils; and (c) toxicity of illicit drugs and their metabolites to biota including daphnia, duckweed, and earthworms. The occurrence and distribution patterns of illicit drugs and their metabolites in specific regions of South Australian wastewater, sewage sludge, surface waters and sediments were investigated. Results indicated that 3 out of 6 illicit drugs were found to be present in wastewaters were at concentrations ranging from 12 to 1670 ng L-1. Methamphetamine (MAP) was the only test compound detected in sewage sludge (2 µg kg-1 dry samples). In surface waters MAP, 3, 4-methylenedioxymethamphetamine (MDMA) and benzoylecgonine (BE) were recorded in 4 out of 20 test locations with a concentration of 5 to 11 ng L-1. Hence, water from wastewater treatment plants (WWTPs) could be the primary source of illicit compounds contaminating the environment. Although the environmental concentrations of these contaminants are low, their impact on aquatic organisms and risk to human health cannot be overlooked. The sorption and desorption patterns of MAP and MDMA (alone and as mixture) were determined in three different soils using batch equilibration experiments. MAP and MDMA reached equilibration within 12 h with initial rapid uptake and then gradually reached equilibrium. Sorption data were analysed employing the Langmuir and Freundlich models, and the results showed that the Freundlich model is the best fit and described the sorption process of MAP and MDMA (alone and as mixture) in three test soils. Sorption of the illicit drugs in soils followed the order: MAP ˃ MAP mixture ˃ MDMA mixture ˃ MDMA. The sorption vii

coefficient (Kd) was positively correlated with soils’ organic carbon (OC), dissolved organic carbon (DOC) and clay for MAP, while for MDMA it was clay, OC and DOC. In addition, the following soil characters such as cation exchange capacity (CEC) ˃ electrical conductivity (EC) ˃ sand were negatively correlated irrespective of the treatments. Furthermore, desorption was assessed when the sorbed particles were released into solution in the following order: MAP ˃ MAP mixture ˃ MDMA ˃ MDMA mixture. These findings could provide an insight into the sorption and desorption patterns, and inform us about the transport and fate of MAP and MDMA in the environment and also their risk assessment. The persistence of cocaine was investigated in a laboratory experiment for 120 days involving three different South Australian soils under both non-sterile and sterile conditions. Cocaine degrades very rapidly in a non-sterile condition in all three test soils (half-life between 2.2 and 3.9 days) compared to sterile condition (half-life between 40.8 and 54.1 days). Cocaine degradation products such as benzoylecogonine (BE) and ecgonine methyl ester (EME) were detected in both conditions. BE was relatively stable for a period of time in non- sterile soil compared to cocaine. Chronic toxicity of illicit drugs (MAP, PSE and cocaine) to earthworm (Eisenia fetida) was studied in a soil. No mortality was recorded even at the highest concentration, and results showed loss in weights for all treatments. Chronic exposure of adult earthworms to MAP, PSE and cocaine showed changes in their morphology and behaviour. Their reproduction capacity also declined especially above 20 mg kg-1 concentration. Exposure at concentrations of 50 – 200 mg kg-1 significantly reduced both cocoon and juvenile stages. Earthworm chronic exposure to cocaine induced and significantly increased DNA damage, olive tail moment, and lipid peroxidation at ˃ 1 mg kg-1 and had a significant impact on total antioxidant capacity at ˃ 25 mg kg-1. Overall, these finding suggests that soil contamination with illicit drugs does constitute a threat to soil biota and the environment The acute and geno-toxicity of MAP, PSE, MDMA and cocaine to a freshwater cladoceran, Daphnia carinata were studied in both cladoceran and natural water collected from local creeks. The cladoceran toxicity followed the order: cocaine ˃ MAP ˃ MDMA ˃ PSE. All these test chemicals were relatively less toxic in non-sterile compared to sterile natural water, which may be due to the influence of varied physico-chemical and biological parameters of natural water. In all the test media, MAP, PSE and MDMA were found to be relatively stable while cocaine was metabolized to BE and ecgonine methyl ester (EME). Also, these chemicals at lower concentrations in water had significant genotoxic effects on D. carinata in comparison to the controls, suggesting that even low level chronic exposure of these compounds to D. carinata can cause serious harm, including developmental and reproductive toxicity.

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The toxic effects of commonly abused illicit drugs and a precursor were assessed by examining their impact on duckweed (Lemna minor L.), a common aquatic plant. Growth attributes (frond numbers, fresh weight and relative growth rate) and biochemical parameters (chlorophyll and proline) content was affected with the increase in the concentrations of illicit drugs. Of these parameters, fresh weight was the most appropriate indicator for validating the effects of illicit drugs. The toxicity of these compounds was followed the order: cocaine ˃ MAP ˃ MDMA ˃ PSE. Overall, the results demonstrate the usefulness of L. minor L. as an illicit drug’s aquatic toxicity indicator for reliable assessment of phytotoxic potential of complex aquatic systems.

Chapter 1 Introduction Illicit drugs are those compounds whose non-medical use is prohibited by international legislation and mainly belong to the classes of opiates, cocaine, cannabis, amphetamines and ecstasy-group substances (UNODC, 2007; Hall et al., 2008). Abuse of these drugs continues to endanger the health and welfare of people which is also a major concern to global economic stability and countries’ social cohesion and developments (UNODC, 2013). According to World Drug Report (2013), these drugs have been used globally at least once in 2012 by 167 to 315 million people (or 3.6% to 6.9% of the world’s population) aged between 15 and 64. The estimated number of drug-related deaths accounted for 2011 was 211,000 and health- related expenses also significantly increase every year. In 2009, cannabis was the most widely used illicit drug, reported to have used up by 125 to 203 million people (i.e. 2.8 to 4.5%) worldwide. In Australia, the illicit drug market has been dominated by cannabis followed by amphetamine-type stimulants (ATSs) in terms of number of drug-related arrests, seizures and use. During 2012-13, nationally, more than 19.6 tonnes of illicit drugs were seized and a record number of seizures (86,918) and arrests (101749) were recorded, which was the second highest reported over the last decade (ACC, 2012-13). In 2011-12, the then highest number of clandestine laboratories (809) was discovered, in which 70% of these were found to be located in residential areas (ACC, 2011-12). Most of them were found to be manufacturing ATSs, because of manufacturing is relatively easier using locally available chemicals and precursors (Pal et al., 2013). Overall, the use of ATSs continued to exceed that of heroin and cocaine combined (UNODC, 2007), and presently several millions of people are reported to be abused by ATSs drugs. Illicit drugs are generally consumed by oral, intranasal (“snorting” powder), by needle injection, or by inhaling smoke (Pal et al., 2013). Following human ingestion, the major portion of the illicit drug’s residues is excreted through urine and faecal matter from human body and released into wastewater networks. The drug residues (parent drugs and its metabolites) may escape degradation and removal processes in wastewater treatment plants (WWTPs) and could become distributed into different environmental compartments (Castiglioni et al., 2007). Moreover, chemical residues and waste materials associated with clandestine drug laboratories are also often covertly disposed into soil, sewage systems, or public waste management facilities (Janusz et al., 2003; Scott et al., 2003). These are considered to be the main source of environmental contamination of illicit drugs. In the environment, illicit compounds tend to undergo various natural processes such as sorption, degradation, leaching, surface runoff, and interactions with different environmental matrices which may

cause these compounds make their way into groundwater and eventually contaminate the environment, thus presenting a major problem for aquatic and terrestrial life including humans (Pal et al., 2013). Our team has studied the degradation, sorption and desorption patterns of certain illicit drugs (parent compounds, precursor, and by-products) in soil and their soil microbiological impacts (Janusz et al., 2003; Scott et al., 2003; Pal et al., 2011, 2012 and 2014). However, information is still lacking on the fate and behaviour of illicit drugs as mixtures of compounds in different soils. The incidence and concentration level of illicit drugs and their metabolites in the environment (e.g., wastewater, surface waters, groundwater, drinking water, and ambient air) and their potential impact on the ecosystem have been mainly reported in European countries, the USA and Canada. In Australia, Irvine et al. (2011) reported the presence of ATSs and BE in wastewaters, but information on surface waters, sediments and sewage sludge is still lacking. Additionally the ecotoxicity of illicit drugs has received less attention compared to pharmaceuticals particularly with reference to low level chronic exposure, possible detrimental effects on aquatic systems, or their bioconcentration in biota (Binelli et al., 2012). These compounds may engage in potent pharmacological activities and their presence as complex mixtures in the environment may cause damage to aquatic and terrestrial organisms. A few studies have reported the ecotoxicity of illicit drugs to aquatic organisms such as zebra fish (Danio rerio), zebra mussel (Dreissena polymorpha) and rainbow trout (Oncorhynchus mykiss) (Pal et al., 2013; Parolini and Binelli, 2013, 2014; Stewart and Kalueff, 2014). Daphnia carinata is one of the most suitable aquatic invertebrate for toxicological research worldwide and it has been well documented in several studies (Cáceres et al., 2007). Specifically, the reasons for using D. carinata are as follows: it naturally occurs in Australian freshwaters; and the size and rate of neonate production makes this species an ideal organism for variety of toxicity studies (Phyu et al., 2004). More importantly, using indigenous taxa to test chemicals’ toxicity does accurately confirm the regional outcomes and eventually minimises variations in testing (Harmon et al., 2003). However, information on the toxicity of illicit drugs and their metabolites to D. carinata is lacking. Furthermore, illicit drugs’ toxicity data on the aquatic plant Lemna minor (Duck weed) is not available. This plant is commonly found in freshwater and is an important food source for various aquatic life forms. L. minor has been used very frequently in ecotoxicological research as a representative of higher aquatic plants and is highly sensitive to different pollutants. This makes L. minor L. a suitable aquatic test organism (Naumann et al., 2007; Drost et al., 2007) to test the illicit drug’s aquatic toxicity. A perusal of available literature shows much less information on the fate of illicit drugs in soil especially their toxicity on soil biota. Soil is currently receiving a substantial amount of

illicit residues from a variety of sources: reused wastewater effluents and sewage sludge containing illicit drugs that end up in crop production and gardening; clandestine laboratories’ waste disposal; landfill leachate; runoff, etc. It must be taken into consideration that emerging contaminants like methamphetamine tend to persist in soil for longer periods (Pal et al., 2011), some illicit drugs (MAP, MDMA and PSE) have a tendency to adsorb to organic carbon (Pal et al., 2014) and notably most of these compounds are biologically active. Therefore, determining the fate and behaviour of illicit drugs in soil environments is essential for continuous monitoring and help us to understand its potential environmental and toxicological threats. To assess the environmental toxicity of chemicals in soil, earthworm (specifically, Eisenia fetida) serves as an ideal organism that has been very often tested for chemicals’ fates in soil. In general, earthworms live in close contact with soil through their skin or gut and process large quantities of soil (Das et al., 2013). Furthermore, E. fetida is very common, easy to handle and culture in laboratory conditions. These ecologically significant characteristics of earthworms make it possible to investigate illicit drugs’ fate when they enter the environment and their potential harmful effects to soil biota. The literature available on the environmental fate and behaviour of the illicit drugs is scanty, given that earlier published work has generally focused on environmental occurrences of illicit drugs and the consequences they have on people’s health. This study has been undertaken with the following objectives in mind: 1. To identify the occurrence and distribution patterns of illicit drugs and metabolites in Australia (SA) in water and soil. 2. To investigate their degradation, and sorption and desorption patterns in soil (alone and as a compound mixture). 3. To study the eco-toxicity of illicit drugs to aquatic and terrestrial biota (daphnia, duckweeds and earthworms).

References ACC (Australian Crime Commission), Illicit drug data report 2011-12. http://www.crimecommission.gov.au/publications/illicit-drug-data-report/illicit-drug- data-report-2011-12 [Accessed on 10th February 2014]. ACC (Australian Crime Commission), Illicit drug data report 2012-13. https://www.crimecommission.gov.au/publications/intelligence-products/illicit-drug- data-report/illicit-drug-data-report-2012-13 [Accessed on July 10, 2014]. Binelli A, Pedriali A, Riva C, Parolini M. Illicit drugs as new environmental pollutants: Cyto- genotoxic effects of cocaine on the biological model Dreissena polymorpha. Chemosphere 2012; 86:906-911. Cáceres T, Megharaj M, Naidu R. Toxicity of fenamiphos and its metabolites to the cladoceran Daphnia carinata: The influence of microbial degradation in natural waters. Chemosphere 2007; 66(7):1264-1269. Castiglioni S, Zuccato E, Chiabrando C, Faneli R, Bagnati R. Detecting illicit drugs and metabolites in wastewater using high performance liquid chromatography-tandem mass spectrometry. Spectro Eur 2007; 19:7-9. Das P, Megharaj M, Naidu R. Perfluorooctane sulfonate release pattern from soils of fire training areas in Australia and its bioaccumulation potential in the earthworm Eisenia fetida. Environ Sci Poll Res 2013; 1-9. Drost W, Matzke M, Backhaus T. Heavy metal toxicity to Lemna minor: studies on the time dependence of growth inhibition and the recovery after exposure. Chemosphere 2007; 67:36-43. Hall W, Degenhardt L, Sindicich N. Illicit drug use and the burden of disease. In: Heggenhougen K, Quah S, editors. International encyclopedia of public health. Elsevier 2008; 523-530. Harmon S, Specht W, Chandler GT. A comparison of the daphnids Ceriodaphnia dubia and Daphnia ambigua for their utilization in routine toxicity testing in the southeastern United States. Arch Environ Contam Toxicol 2003; 45(1):79-85. Irvine RJ, Kostakis C, Felgate PD, Jaehne EJ, Chen C, White JM. Population drug use in Australia: a wastewater analysis. Forensic Sci Inter 2011; 210:69-73. Janusz A, Kirkbride KP, Scott TL, Naidu R, Perkins MV, Megharaj M. Microbial degradation of illicit drugs, their precursors, and manufacturing by-products: implications for clandestine drug laboratory investigation and environmental assessment. Forensic Sci Inter 2003; 134:62-71.

Naumann B, Eberius M, Appenroth KJ. Growth rate based dose-response relationships and EC-values of ten heavy metals using the duckweed growth inhibition test (ISO 20079) with Lemna minor L. clone St. J Plant Physiol 2007; 164:1656-1664. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor and by-products in soil. Chemosphere 2011; 85:1002-9. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Illicit drugs and the environment—A review. Sci Total Environ 2013; 463-464:1079-1092. Pal R, Megharaj M, Kirkbride KP, Naidu R. Adsorption and desorption characteristics of methamphetamine, 3, 4-methylenedioxymethamphetamine, and pseudoephedrine in soils. Environ Sci Poll Res 2014; 1-11. DOI 10.1007/s11356-014-2940-6. Pal R, Megharaj M, Kirkbride KP, Naidu R. Fate of 1-(1′, 4′-cyclohexadienyl)-2- methylaminopropane (CMP) in soil: Route-specific by-product in the clandestine manufacture of methamphetamine. Sci Total Environ 2012; 416:394-399. Parolini M, Binelli A. Adverse effects induced by ecgonine methyl ester to the zebra mussel: A comparison with the benzoylecogonine. Environ Pollut 2013; 182:371-378. Parolini M, Binelli A. Oxidative and genetic responses induced by Δ-9-tetrahydrocannabinol (Δ-9-THC) to (Dreissena polymorpha). Sci Total Environ 2014; 468:68-76. Phyu YL, Warne MSJ, Lim R. Toxicity of atrazine and molinate to the cladoceran Daphnia carinata and the effect of river water and bottom sediment on their bioavailability. Arch Environ Contam Toxicol 2004; 46(3):308-315. Scott TL, Janusz A, Perkins MV, Megharaj M, Naidu R, Kirkbride KP. Effect of amphetamine precursors and by-products on soil enzymes of two urban soils. Bull Environ Contam Toxicol 2003; 70:824-31. Stewart AM, Kalueff AV. The behavioral effects of acute Δ9-tetrahydrocannabinol and heroin (diacetylmorphine) exposure in adult zebrafish. Brain Res 2014; 1543(0):109-119. UNODC (United Nations Office on Drugs and Crime). World drug report. United Nations Publication; 2007. http://www.unodc.org/pdf/research/wdr07/WDR_2007.pdf [Accessed on February 25, 2014]. UNODC (United Nations Office on Drugs and Crime). World Drug Report. United NationsPublication.2013.http://www.unodc.org/unodc/secured/wdr/wdr2013/World_D rug_Report_2013.pdf . [Accessed on February 25, 2014].

Chapter 2

LITERATURE REVIEW Illicit drugs environmental occurrence, fate and toxicity

Abstract Illicit drugs are the latest group of emerging pollutants and determining their concentrations in wastewater, surface waters, sewage sludge and atmospheric air is an indirect method for estimating the community level of consumption. These substances enter the wastewater systems as unchanged parent drugs and/or their metabolites by human excretion after illegal consumption or by accidental or deliberate disposal from clandestine drug laboratories. Despite the current environmental concentrations of illicit drugs being low, cocaine, morphine, amphetamine, and MDMA possess potent pharmacological activities and their presence as complex mixtures in water may adversely affect aquatic organisms and risk human health. The fate of illicit drugs in water streams depends on various factors such as compounds’ physico-chemical properties, seasonal variations (temperature and rain), and the techniques used for removing illicit drugs in wastewater treatment plants. We reviewed the literature on illicit drugs and their metabolites in different environmental scenarios (e.g., wastewater, surface waters, groundwater, drinking water, sewage sludge and ambient air), their fate in water and soils and their toxic effects on aquatic and terrestrial biota. Illicit drugs occur in different environmental contexts, and aquatic and terrestrial organisms may be affected by these compounds. Further research is required on the concentrations of illicit drugs with reference to manufacture and high use, the environmental fate of these compounds, and their impact on ecosystems at concentrations that typically occur in the environment.

2.1 Introduction Drug abuse has become a serious international problem due to substances’ increasingly harmful effects on those who produce and consume them, and to the environment (Pal et al., 2013; Rieckermann and Christakos, 2008). Illicit drugs are prohibited by national and international laws because they have non-medical uses which are plant and synthetic derivatives (Hall et al., 2008). These drugs include amphetamine type stimulants (ATSs), cannabis, cocaine, heroin and other opioids. Globally, in 2013, 167 to 315 million people used any of these drugs at least once, and their ages ranged between 15 and 64. Cannabis is the most widely used illicit drug type and it is consumed by 120 to 224 million people worldwide. The drugs that follows cannabis are ATSs; mainly methamphetamine (MAP), amphetamine,

and ecstasy, opiates (opium, heroin and prescription opioids) and cocaine (UNODC, 2013). In Australia, ATSs is the second most used illicit drug after cannabis and 809 clandestine laboratories have been detected in recent years (ACC, 2011-12). The plant-based illicit drugs opium and cocain are produced in certain parts of the world and transported to other countries (UNODC, 2010; Mejía and Restrepo, 2008; Schiff, 2002). Cannabis plants can be grown indoors (hydroponics) or in a controlled condition for higher delta-9-terrahydrocannabinol (THC) concentrations (McLaren et al., 2008). These plant derivatives’ production and consumption have entered a stable global market phase in recent years 2005-2006 (UNODC, 2008b, 2007). However, the ATSs global market is continuing to grow in terms of manufacture and consumption (UNODC, 2008a), and currently, 53.2 million people are reported to be users of ATSs (UNODC, 2013). In reality, abuse of these psychoactive substances is almost universal (Pal et al., 2013). There are a number of reasons for the ATSs global market expansion such as their manufacture requires minimal effort and are easy to hide, need only minimal space to set up, have easily available precursor chemicals, and create high psychoactive outcomes for users (UNODC, 2008a). In the past few years, illicit drugs and their metabolites have been identified worldwide from different environmental systems, and they have become a latest group of emerging pollutants (Castiglioni and Zuccato, 2010; Pal et al., 2013). Illicit drugs and their metabolites pollute the environment mainly through human consumption and clandestine waste disposal. In the human body the major portion of administered pharmaceuticals and illicit drugs are metabolised and excreted via urine and faeces as parent compounds and/or metabolites, and discharged directly into the sewage system (Al-Rifai et al., 2011; Castiglioni et al., 2007; Zuccato et al., 2005). Clandestine waste materials such as parent compounds, precursors and by-products of illicit drugs are also disposed of illegally into water bodies, the soil, toilets and eventually returning to the environment via wastewater recycling (Janusz et al., 2003; Scott et al., 2003). Some of these illicit compounds are highly polar and persist in the environment which may pose a greater threat to wildlife and humans. These illicit compounds may have potent pharmacological activities and their presence in water bodies even at low concentrations, together with the residues of several related compounds may prove dangerously toxic to aquatic organisms. It is clear that illicit drugs’ environmental, wildlife and human health-related issues cannot be easily ignored (Castiglioni et al., 2007; Pal et al., 2013; Pomati et al., 2006; Zuccato et al., 2008). Analysis of illicit drugs in water bodies via mass spectrometry, and their environmental occurrence and distribution patterns in the environment have been reviewed, respectively, by several researchers (Castiglioni et al., 2008; Pal et al., 2013; Postigo et al., 2008a, b). However, comprehensive information on illicit drugs’ occurrence in the environment, their fate,

and toxicity to aquatic biota and mammals has not yet been available. We must have more knowledge on illicit drugs’ environmental pathways and factors controlling their environmental fate and toxicity, so that we can develop an essential tool to combat these drugs’ environmental contaminations and associated health problems. Consequently, the aim of this review is to critically review the occurrence, fate and impact (ecotoxicity) of illicit drugs in different environmental systems.

2.2 Environmental occurrence of illicit drugs and their metabolites Illicit drugscontamination of the environment is occurring mainly due to legal or illegal human consumption and/or by manufacturers’ disposing of them in sewage systems (González-Mariño et al., 2012; Boles and Wells, 2010; Pal et al., 2013). Illicit drugs can be ingested by people in different ways - injecting, smoking, snorting, and swallowing (Pal et al., 2013; Verster, 2010; Vazquez-Roig et al., 2010). After human consumption these substances are excreted as parent compounds and/or as metabolites in urine, faeces, saliva, and sweat, and the proportion of excretion varies with the drugs (Table 2.1) and the individual drug user’s body state (van Nuijs et al., 2011a, b; Boles and Wells, 2010; Gheorghe et al., 2008; Chiaia et al., 2008). Illicit drugs’ residues found in users’ waste products (urine and faeces) that enter the sewage system with wastewater are only partially removed by wastewater treatment plants (WWTPs) (Zuccato and Castiglioni, 2009). The efficiency in removing illicit compounds from effluent waters depends on: firstly, what technologies WWTPs employ; and secondly, the nature of the compounds (Rodayan et al., 2014; Zuccato and Castiglioni, 2009; Bijlsma et al., 2009; Huerta-Fontela et al., 2008a). The removal efficiencies of commonly abused drugs are summarized in Table 2.1. Overall, efficiency in removing illicit drugs has been reported in the following order in terms of drug groups: ATSs ˃ cannabinoids ˃ opioids ˃ cocainics (Postigo et al., 2010). These substances have consequently been identified using different environmental systems (wastewater, surface water, tap water, sewage sludge and atmospheric air) in the high ng L-1 to low µg L-1 range (Pal et al., 2013).

Table 2.1 Selected physico- chemical properties of major illicit drugs and metabolites

Drugs Molecular Molecular Precursors (d) Human metabolic by- Removal Log KOW, weight structure products and excretion efficiency in PKa rate (c) WWTP (%) (e) (f) Plant based drugs Cocaine 303.36 Dried coco leaves, Cocaine 1 - 9% 72 - 100% 2.30, 8.6 (C17H21NO4) atropine, tropinone, Benzoylecogonine 35 - carbomethoxytropinone 54% Ecgonine methyl ester 32 - 49%

Benzoylecogonine 289.33 - - 83 - 100% -1.32, 2.15 (C16H19NO4)

Ecgonine methyl 199.25 - - - ester (C10H17NO3)

Morphine 285.34 Opium extracts, Morphine 6.8% b 72 - 98% -0.1, 8.0 (C17H19NO3) codeine Conjugated morphine 58.6% b

Heroine 369.42 morphine, diamorphine, Heroine 0.1% - 1.58a, 7.95a (C21H23NO5) acetic anhydride Morphine 4.2% Conjugated morphine (mainly M3G & M6G) 38.3%

9 9 a a Δ -THC (C21H30O2) 314.47 - Δ -THC 10-25%, 72 - 99% 6.97 , 10.6 11-hydroxy-THC (OH-THC) 2%, 11-nor-9-carboxy-THC (THCCOOH) 0.5% Synthetic drugs (ATS and ecstasy drugs) Amphetamine 135.21 1-phenyl-2-propanone, Amphetamine 30% 52 - 99% 1.76, 10.1 ((C9H13N) norephedrine, Phenylacetone 0.9% norpseudoephedrine Hippuric acid 16 - 28%

Methamphetamine 149.23 Ephedrine, Methamphetamine 43% 44 - 99% 2.07, 9.9 (C10H15N) pseudoephedrine, Amphetamine 4 - 7% 1-phenyl-2-propanone p-hydroxy methamphetamine 15% MDMA 193.25 safrole, isosafrole, MDMA 26 - 65% 44 - 57% 2.28, NA (C11H15NO2) piperonal, 3,4- MDA 1% methylenedioxyphenyl-2- propanone MDA 179.22 - - 60% 1.64, NA (C10H13NO2)

References: (a) http://pubchem.ncbi.nlm.nih.gov/compound/Diacetylmorphine (b) Mitchell et al. (1991) (c) Postigo et al. (2008a) (d) http://www.emcdda.europa.eu/drug-profiles (e) Pal et al. (2013)

2.2.1 Wastewater Extensive studies have demonstrated presence of various pharmaceuticals and personal care products (PPCPs) in aquatic ecosystems (Rosi-Marshall et al., 2014). Recently, several studies have investigated the low level occurrence (i.e., ng L-1) of illicit drugs’ parent compounds and their metabolites in water bodies (Pal et al., 2013). Most analyses have focused almost exclusively on urban or moderately-sized rural wastewater treatment plants, often investigating treatment efficacy for specific compounds (Rosi-Marshall et al., 2014). The occurrence of illicit drugs in aquatic environments is of significant interest to environmental scientists (Pal et al., 2013). Detecting illicit drugs’ residues in wastewater could provide evidence-based real-time data on type and level of community-wide illicit drug use (Castiglioni et al., 2008; Bones et al., 2007). Monitoring illicit drugs in wastewater not only provides government authorities with data on the nature and magnitude of drug abuse, but also reveals the drug abuse trend in a particular area (Castiglioni et al., 2014; Rieckermann and Christakos, 2008; Terzic et al., 2010). This may help environmental scientists and policy-makers to implement control strategies to protect the environment from these pharmacologically active substances (Pal et al., 2013). Compared to PPCs, information regarding illicit drugs’ existence in the environment and particularly aquatic systems is relatively limited. In recent years, the number of studies reporting the presence of illicit drugs in wastewaters has grown: Australia (Irvine et al., 2011; Lai et al., 2011); Belgium (Gheorghe et al., 2008; van Nuijs et al., 2009a, b, c, d, 2011a, b); Canada (Rodayan et al., 2014; Metcalfe et al., 2010); Croatia (Terzic et al., 2010); France (Damien et al., 2014; Karolak et al., 2010); Germany (Hummel et al., 2006); Ireland (Bones et al., 2007); Italy (Repice et al., 2013; Castiglioni et al., 2006, 2007; Zuccato et al., 2005; Mari et al., 2009); the United States (Jones-Lepp et al., 2004; Chiaia et al., 2008; Loganathan et al., 2009; Bartelt-Hunt et al., 2009); Slovakia (Mackuľak et al., 2014); Spain (Bijlsma et al., 2014; Vazquez-Roig et al., 2014; Huerta-Fontela et al., 2008a, 2007; Boleda et al., 2007, 2009; Postigo et al., 2008b, 2010, 2011; Bijlsma et al., 2009; González-Mariño et al., 2010, 2012; Bueno et al., 2011; Pedrouzo et al., 2011); Switzerland (Berset et al., 2010); and the UK (Baker et al., 2014; Kasprzyk-Hordern et al., 2008, 2009, 2010). The commonly identified parent compounds and their metabolites in wastewater include: cocaine, benzoylecgonine (BE), norbenzoylecgonine, norcocaine, cocaethylene, and ecgonine methyl ester (EME) (van Nuijs et al., 2011a); opioids (morphine, 6-acetylmorphine, morphine-3β-D-glucuronide, methadone, and 2-ethylidine-1,5-dimethyl-3,3- diphenylpyrrolidine (EDDP)) (Castiglioni et al., 2006, 2007; Postigo et al., 2008b, 2010; Bueno et al., 2011; Terzic et al., 2010); cannabinoids (Δ9-tetrahydrocannabinol (THC)) (Castiglioni et al., 2006, 2007; Postigo et al., 2010; Boleda et al., 2007 and 2009); amphetamine and ecstasy

group compounds (amphetamine, MAP), 3,4- methylenedioxy methamphetamine (MDMA), 3,4-methylenedioxy amphetamine (MDA), 3,4-methylenedioxyethylamphetamine (MDEA)) (Zuccato et al., 2005; Castiglioni et al., 2006; Postigo et al., 2008b; Gheorghe et al., 2008). The reported concentrations (ng L-1) of cocaine and its metabolites, opioids and cannabinoids in wastewater influent and effluent ranges between 18.8 – 7500 and 7.5 – 3425; while the concentrations of ATSs recorded 2.2 – 15380 and 1.0 - 10955 (Table 2.2) (Pal et al., 2013). According to the available literature, higher concentrations of cocaine and its metabolites exist in the WWTPs influent in Spain, Italy, and Switzerland, while the lowest concentrations were documented for France, the USA, and Australia (Pal et al., 2013). Most reports indicated that the higher concentrations of the metabolite are comparable to the parent compound which agrees with the known human metabolism of cocaine (Zuccato et al., 2005). Higher concentrations of amphetamine have been documented in the British and Spanish WWTPs’ influent while MAP from the USA, and MDMA from Spain (Pal et al., 2013). Generally, the illicit drugs and their metabolites concentrations in effluent were below the limit of quantification (LOQ) to very low ng L-1, which depended on the nature of the drugs and the technologies used in WWTPs (Boles and Wells, 2010; Zuccato and Castiglioni, 2009; Huerta- Fontela et al., 2008a). Illicit drugs’ concentrations in the influents indicate possible drug use pattern in the local community, while that in effluent reflects the potential for the contamination of the receiving water bodies (Huerta-Fontela et al., 2008a; Zuccato et al., 2008). These illicit residues can be moved or transported to other places via run-off during wet weather, landfill leachate, aquifer recharge, and farm land drainage where effluent and sewage sludge are used for crop production (Boles and Wells, 2010; Jones-Lepp et al., 2004; Kaleta et al., 2006).

2.2.2 Surface waters Illicit drugs contamination of surface waters is primarily due to the wastewater effluents from WWTPs (Pal et al., 2013). Illicit drugs reach the surface water as unaltered parent compounds or transformed metabolites (Boleda et al., 2009, 2011). Existing wastewater treatment processes do not completely remove these illicit substances from the wastewater (Valcárcel et al., 2012). The efficiency in removing illicit drugs from the wastewater varies significantly according to their physico-chemical properties, climatic factors (temperature, rainfall, sunlight, etc.) and technology being used such as advanced oxidization process, ozonation, osmosis, etc. (Terzic et al., 2010; Kasprzyk-Hordern et al., 2009). Consequently, illicit substances are detected in water bodies including rivers, lakes, and groundwater, and this is mainly because compounds have not been completely removed during wastewater treatment and/or by discharge of manufacturing residues (Roberts and Thomas, 2006; Al-Rifai et al., 2007).

Several studies have reported the presence of these substances in surface water. The observed concentrations of cocainics, opioids and cannabinoids were between 0.1 and 319 ng L-1 while the ATSs were between 0.4 and 309 ng L-1 (Table 2.2) (Pal et al., 2013; Zuccato et al., 2005, 2008; Kasprzyk-Hordern et al., 2008, 2009; Huerta-Fontela et al., 2008b, 2007; Boleda et al., 2007, 2009; Vazquez-Roig et al., 2010; Postigo et al., 2010; González-Mariño et al., 2010; Bueno et al., 2011; Valcárcel et al., 2012; Gheorghe et al., 2008; van Nuijs et al., 2009c, d; Berset et al., 2010). A relatively higher concentration of cocaine and BE was reported in Spanish, Belgian, and Italian surface waters, while the amphetamine and MDMA were recorded as being higher in Spain; MAP was higher in the USA (Pal et al., 2013; Zuccato et al., 2008; González-Mariño et al., 2010; Bueno et al., 2011; Huerta-Fontela et al., 2008b; Bartelt-Hunt et al., 2009). Recent studies have revealed the presence of illicit drugs and their metabolites even in drinking water treatment plants (Huerta-Fontela et al., 2008b; Boleda et al., 2009, 2011; Valcárcel et al., 2012) and in groundwater (Jurado et al., 2012). In addition, cocaine, BE, MDMA, methadone and its metabolite EDDP were frequently detected in tap water as well as in groundwater (Boleda et al., 2011; Valcárcel et al., 2012; Jurado et al., 2012).

2.2.3 Sewage sludge and bio - solids The presence of illicit drugs in sewage sludge, bio-solids, and sediments has received little attention compared to wastewater and surface waters (Pal et al., 2013). Kaleta et al. (2006) were the first to report the presence of amphetamine (5 - 300 μg kg−1) in sewage sludge (Austria) while in the USA, Jones-Lepp and Stevens (2007) reported MAP (4 μg kg−1 dry weight) in bio-solids. Wick et al. (2009) recorded the sorption coefficient (Kd) at 12 and 76 L kg-1 for morphine and methadone, respectively, in sediments. Recently, Mastroianni et al. (2013) reported (Spain) the presence of cocaethylene, ephedrine, heroin, alprazolam, lysergic acid diethylamide (LSD), its metabolite 2-oxo-3-hydroxy-LSD, and the cannabinoids (THC), cannabinol (CBN) and cannabidiol (CBD) in sewage sludge. Furthermore, the cannabinoids, methadone and its metabolite EDDP were the most ubiquitous and abundant compounds in sewage sludge, where the concentration recorded more than 100 ng g-1 dry weight in all target compounds. At the same time THC was up to 579 ng g-1 dry weight.

2.2.4 Atmospheric air Some illicit drugs such as cocaine, cannabis, heroine and MAP have the potential to escape into the atmosphere during consumption and/or handling (Daughton, 2011; Viana et al., 2010; Pal et al., 2013). Illicit drugs and their metabolites in the ambient air can be easily detected and quantified because of their physico-chemical properties such as low vapour

pressures, high to medium polarity, weak alkalinity, and molecular weight (Pal et al., 2013; Postigo et al., 2009; Viana et al., 2010). Not much information is currently available on the presence of illicit drugs in ambient air. To date only a few countries (Italy, Spain, Portugal, Chile, Brazil, Serbia, and Algeria) have reported the presence of these substances in ambient air. According to the data, cocaine was detected maximum concentration in the atmospheric air (low ng m−3) followed by THC (low pg m−3) while the synthetic illicit drugs were very low (Balducci et al., 2009; Cecinato et al., 2009a, 2009b; Viana et al., 2010, 2011).

2.3 Environmental fate of illicit drugs 2.3.1 Water Illicit drugs’ parent chemicals are often excreted from the human body with a number of associated metabolites. For example, cocaine is excreted as unchanged (1 - 9%) and several metabolites: BE (35 - 54%), EME (32 - 49%), cocaethylene (0.7%) and ecgonine (Postigo et al., 2008a, b). Major drugs that people abuse and their metabolic by-products are summarized in Table 2.1. Wastewater treatment processes play an important role in the life cycle of illicit drugs in water systems (Evgenidou et al., 2015). The main transport pathways of these compounds into the environment are via treated wastewater reuse where they may be only partially removed (Pal et al., 2013). Efficiency in removing illicit drugs during the water treatment process depends on the methods employed in a specific water treatment plant and contaminants’ physico-chemical properties, such as degradation potential, water solubility, adsorption and tendency to volatilize (Behera et al., 2011; Luo et al., 2014; Evgenidou et al., 2015). Of these physico-chemical properties the adsorption coefficient of a compound is an important variable that affects wastewater treatments’ removal capacity. The sorption process of a compound facilitates an interaction with the solid particles (settling, flotation) or biological processes (biodegradation). Subsequently, compounds with low adsorption coefficients tend to remain in the aqueous phase, which favours their mobility through the treated wastewater and into the receiving waters (Ohlenbusch et al., 2000). Furthermore, compounds with lower log Kow values (˂ 3.0) are not expected to be adsorbed significantly to the particles, and therefore they exhibit low removal efficiency in the water treatment (Behera et al., 2011). According to the available literature, the commonly abused illicit drugs (morphine, amphetamine, MAP, MDMA, MDA and cocaine) have log Kow values between -0.1 and 2.39 (Rosi-Marshall et al., 2014). At the same time, contaminants with relatively high log Kow values and pKa values below the pH of the wastewater are likely to be dissociated in the aqueous phase and not bound to the particles (Thomas and Foster, 2005). Illicit drugs in the aquatic system tend to degrade which leads to the formation of various transformation products (TPs). These TPs can be equally or even more toxic and

dangerous than the parent compounds. Some parent drugs can resist microbial degradation where abiotic degradation is likely to occur in wastewater involving direct or indirect photodegradation, hydrolysis and oxidative reactions (Evgenidou et al., 2015). Cocaine degradation in wastewater depends on pH, temperature and time, and in one study the rate of degradation was minimal at -20°C and pH 2 but it was substantial at +20°C and pH 6 (Gheorghe et al., 2008). Gheorghe et al. (2008) and Postigo et al. (2008b) have reported that cocaine was degraded very quickly in surface waters (75% after 1 day) at 20°C compared to HPLC grade water, which could have been due to the microbial degradation process and other processes. Cocaine degradation was associated with transformation into its metabolites such BE and EME (Gheorghe et al., 2008; Postigo et al., 2008a). Recent studies on illicit drugs’ stability in water revealed that ATSs (amphetamine, MAP and MDMA) were stable for a period of time compared to other illicit drugs (Thai et al., 2014; van Nuijs et al., 2012). Amphetamine and MAP are photo-stable as well, and their degradation occurred in a river water predominantly through microbial activity (Bagnall et al., 2013). Overall, the available data emphasizes that the fate of illicit drugs in water bodies depends on various factors such as compounds’ properties, seasonal variations, microbial activity and techniques used to remove illicit drugs in wastewater treatment plants. Reports have not yet been published on experiments for the degradation of LSD (lysergic acid diethylamide) and it metabolites, heroine, nor-morphine, THC and its metabolic by-product in water (Postigo et al., 2008a).

2.3.2 Soil/sediments/sludge Until now only a few studies have investigated the fate and behavior of illicit drugs in soils, sewage sludge and sediments. For example, the partitioning of illicit drugs in sediments and sewage sludge has been examined for a few drugs such as morphine, cocaine, BE and EME (Stein et al., 2008; Plósz et al., 2013). In a kinetic experiment Stein et al. (2008) studied the sorption of morphine in two sediments and found significant dissipation of morphine, even in the presence of sodium azide, which complicates the interpretation of morphine sorption kinetics. The observed log Koc values for morphine in the two sediments ranged from 2.6 to 2.7. Plósz et al. (2013) evaluated the sorption of cocaine and its metabolites (BE and EME) in sewage sludge at neutral pH, and they determined the linear partition coefficients (Kd) equal to 8400; 200; and 300 L kg-1 for cocaine, BE and EME, respectively. Generally the sorption of illicit drugs is likely to be pH-dependent, as several illicit drugs will be charged at pH ≥ 8 and many illicit drugs are highly polar and may exhibit hydrogen bonding (Stein et al., 2008). Pal et al. (2014) investigated the sorption and desorption of MAP, MDMA and PSE in a batch

equilibrium experiment for three different soils, where they found Koc values between ˂ 250 to ˂ 400 for MAP, MDMA and PSE in most cases. In fact, the Koc value has implications for the leaching potential of organic chemicals and especially their potential risk for groundwater contamination. Wadaskar et al. (2006) noted that organic chemicals with Koc values < 50 may be considered to be highly mobile, those with values of 150 - 500 to be moderately mobile, and those with Koc > 2,000 to be only slightly mobile compounds. Furthermore, Pal et al. (2014) described the following order of sorption (MDMA ˃ MAP ˃ PSE) and desorption (PSE ˃ MAP ˃ MDMA) pattern of illicit drugs in three different soils. Overall, the sorption and desorption of MAP, MDMA, and PSE in soil mainly depended on the particular solute chemical structure and the adsorbent properties of soil constituents (Pal et al., 2014). In spite of having only limited data on the fate and behavior of illicit drugs in soils, the persistence of MAP, MDMA and PSE has been reported by Pal et al. (2011). They summarized the following descending order of half-life in non-sterile soil as follows: MAP (131 to 502 days) ˃ MDMA (15 to 59 days) ˃ PSE (3.7 to 30 days). Generally, the reported half- lives were longer in sterile conditions compared to non-sterile soils, suggesting that microbially-mediated biotransformation of illicit drugs can be significant in soils (Pal et al., 2011). The degradation pattern of methylamphetamine sulfate (MAS) and its precursor P2P (phenyl-2-propanone), and manufacturing by-product in soil was initially reported by our research group (Janusz et al., 2003). In their study, Janusz et al. (2003) observed a rapid degradation of P2P in all the tests soils while the degradation of MAS was very slow with the level remaining practically constant over a period of 6 weeks. As very little information is available on the fate and behavior of illicit drugs in solid matrices, it is difficult to draw a conclusion and in future it will be important to further evaluate the sorption, desorption and persistence of commonly abused illicit drugs individually and as a mixture of compounds in a varied solid biomass (sludge and sediments) and soils.

2.4 Toxicity of illicit drugs Generally illicit drugs’ exposure is harmful to humans, aquatic and terrestrial biota. Several studies have been done on the presence of illicit drugs in different environmental systems worldwide (Pal et al., 2013). However, the research focus on illicit drugs’ ecotoxicity has received less attention compared to legal pharmaceuticals, in particular their potential direct and indirect toxicity on non-target aquatic systems and their associated biota (Daughton, 2011; Binelli et al., 2012). The principal sources of illicit drugs in the environment are drug residues excreted after illegal human consumption, manufacturers’ disposal of waste into the domestic sewage system or leaky landfills (Pal et al., 2013). Though the reported illicit drugs’ environmental concentrations are low (cocainics between 0.1 and 7500 ng L-1 and ATSs

between 0.2 - 15380 ng L-1), the outcomes of chronic exposure to aquatic and terrestrial biota including humans cannot be ignored (Parolini et al., 2013). In fact, these psychoactive substances are biologically active, and their presence as complex mixtures along with other chemicals in water bodies, may lead to unforeseeable pharmacological interactions that are harmful to aquatic organisms (Pomati et al., 2006; Parolini et al., 2013).

2.4.1 Toxic effects on aquatic organisms Aquatic environments are the main destination for a major portion of illicit compounds’ residues produced by humans (Binelli et al., 2012). However, in comparison to other environmental pollutants, relatively little is known about the illicit drugs’ toxicity on aquatic organisms (Binelli et al., 2012; Parolini et al., 2013). To date, there are only a few studies on the ecotoxicity of cocaine, BE, EME, amphetamine, and morphine on aquatic organisms (Table 2.3). Cocaine is abused worldwide, and as a result its residues have been recorded in different environmental systems (Pal et al., 2013; UNODC, 2013). Cocaine has been tested for its behavioral alteration and sensitivity to mutagenized Zebra fish (Danio renio), in which higher cocaine concentration reduced the visual sensitivity and altered the dopaminergic signalling in the brain as a result of mutation in distinctive genes (Darland and Dowling, 2001). A study on the cyto-genotoxic effects of cocaine on Zebra mussel (Dreissena polymorpha) indicated: significant DNA damage; increase in micro nucleated cells; and noticeable rise in apoptosis (Binelli et al., 2012). Cocaine was also found to cause the mitotic stimulation process in the protozoan Tetrahymena pyriformis, as a result of the following stimulation of the DNA synthesis. This indicates that cocaine products (salt and freebase) have a mitogenic effect (Stefanidou et al., 2002). Parolini et al. (2013) reported that BE at a higher concentration (1 µg L-1) induced a oxidative stress in D. polymorpha, in which significant increases of lipid peroxidation (LPO) and protein carbonyl content (PCC) occurred, besides primary and fixed DNA damage. At the same time it was observed that both tested concentrations (0.5 & 1µg L-1) were involved in the oxidative stress. This was due to BE toxicity which compromised lysosomal membrane stability and imbalanced defence enzyme activities (Parolini et al., 2013). In a another study on D. polymorpha, BE significantly altered important metabolic proteins in gill cells, which is believed to be a possible effect on calcium homeostasis and a subsequent imbalance of oxidative stress (Binelli et al., 2013). Cocaine’s second major metabolite EME also has the ability to cause serious problems for D. polymorpha, such as weakening of the lysosome membranes, complete inactivation of defence enzymes, increases in LPO, protein carbonylation, and DNA fragmentation. However,

there was no change in fixed genetic damage to D. polymorpha such as that with BE. Simultaneously, both the cocaine metabolites (BE and EME) in biomarker response index at 0.5 mg L-1 indicated equivalent toxic effects on D. polymorpha (Parolini and Binelli, 2013). In addition, chronic cocaine exposed European Eel (Anguilla anguilla) showed a significant tissue accumulation of cocaine at a specific environmental concentration (20 ng L-1) (Capaldo et al., 2012). Since the A. anguilla is a less active fish with high fat content and making it susceptible to a significant bioconcentration of cocaine, this may pose a risk to fish health. For instance, reproduction could be reduced by mediating dopamine receptors (Nestler and Malenka, 2004; Sébert et al., 2008; Capaldo et al., 2012). Amphetamine sulfate is a psychoactive stimulant which was first tested for its toxicity to freshly isolated rainbow trout’s (Oncorhynchus mykiss) hepatocytes and Daphnia magna, and the results show that of the 50 reference chemicals used, amphetamine sulfate was highly cytotoxic (Lilius et al., 1994). A morphine pharmacokinetics study revealed a significant intra- specific variation of plasma morphine concentrations between winter flounder (Pseudopleuronectes americanus) and sea water acclimated rainbow trout (Oncorhynchus mykiss). The disposition of morphine in fish was less than in mammals. This variation may be associated with the mass specific differences in cardiac output (Newby et al., 2006). Furthermore, morphine indicated immunotoxic effects on freshwater mussel (Elliptio complanata) where morphine significantly reduced the phagocytic and intracellular esterase activity, cell adherence, and LPO in the hemolymph (Gagné et al., 2006). Recently, Parolini and Binelli. (2014) have documented that 0.5 µg L-1 of THC on D. polymorpha induced LPO, protein carbonylation and DNA damage due to oxidative stress. In addition, the same concentration of THC exhibited some toxicity on aquatic species then BE and EME, based on BRI values, which suggests that THC is a threat to aquatic organisms (Parolini and Binelli, 2014). This data can also be used to evaluate and predict the toxicity of illicit drugs on humans and mammals. Generally, most chemicals exercise their influence through cellular functions which is common in both category cells (Ekwall, 1983). Recently, Stewart and Kalueff (2014) observed the THC and heroin-induced hypolocomotor and hyperlocomotion effects on zebra fish.

2.4.2 Toxic effects on animals 2.4.2.1 Cocaine Cocaine can be toxic to animals through various mechanisms (Tseng et al., 1993). It produces its major effects by blocking the reuptake of catecholamine like dopamine, serotonin and norepinephrine into presynaptic terminals, which play an important role in removing the neurotransmitters from the extracellular space leads to activate receptors (Cunningham et al., 18

1995; Izenwasser, 1998). Some findings are contrary to that cocaine is not causing catecholamine release or alteration (McKenna and Ho, 1980). However, dopamine uptake is the primary mechanism of cocaine in influencing the behaviour of animals, based on the exposure level (Izenwasser, 1998). The effects of cocaine on animals varied with exposure period (acute and chronic) and concentration; the observed effects ranged from mild impact to death or seizure (Tseng et al., 1993; Macêdo et al., 2005; Numa et al., 2008). In low dose (1 mg kg-1) cocaine exposure induced aggression in rats compared to control and high dose (20 mg kg-1) while in mice cocaine did not induce any aggressive behavior in both single and group situations, irrespective of the concentrations tested. Instead, it led to reduced social contacts which may be associated with cocaine anxiogenic effects (Estelles et al., 2004). Cocaine can induce temporary acute anxiolytic-like behaviour after first exposure and this rapidly led to the anxiogenic state in rodents (Müller et al., 2008). Like cocaine, coca-paste (CP) wields acute anti-aggression influence on male rats, which may be associated with the effect of accumbal dopamine and cortical serotonin level (Meikel et al., 2013). Rats displaying anxiety are highly receptive to cocaine exposure and this may increase the chances of being addicted to cocaine, compared to non-anxious rats (Pelloux et al., 2009). Prenatal exposure to cocaine could create several behavioral changes during aging in rats (Sobrian et al., 2003) which was evident in low dose of cocaine exposed pre-weanling rats (18 -21 days). They showed distinguishable differences during the adult stage, such as being active for longer periods, rapid activity peaks and the development of less tolerance patterns. This could be due to the CNS having longer sensitivity in rats’ early development stage to stimulant drugs (Smith and Morrell, 2008). Acute cocaine administration produces a time-dependent memory deficit in rodents which occurred after termination of the locomotors’ stimulant effect of cocaine (Niigaki et al., 2010), and in some cases cocaine’s residual effects on working and long-term memory deficit in preadolescent rats was noticed (Morrow et al., 2002). However, it may be recoverable as the animal ages (Santucci et al., 2004). Similar to adults the in utero cocaine (IUC) exposure caused significant learning impairments and memory deficits in juvenile male and female offspring rats due to neurodevelopmental damage (Melnick et al., 2001; Meunier and Maurice, 2004). Adding to the effect on animals’ behavior, apoptosis activity was noticed in the cortical neurons of fetal mice (Nassogne et al., 1997). Acute and/or chronic exposure of cocaine induced apoptosis in fetal rat myocardial cells (Xiao et al., 2000), and the proapoptotic genes expressions increased in developing cerebral cells (Lee et al., 2009). However, Binelli et al. (2012) believe that the behavioral changes in animals were associated with induction of apoptosis or necrosis activity. Furthermore, exposure to cocaine revealed an increase in

apoptosis activity in rat hepatocytes, which was correlated to elevated oxidative stress (Díez- Fernández et al., 1999). In most situations, cocaine is abused along with other substances, for example caffeine. The combination of caffeine + cocaine reported in the canine model produced significant synergistic chronotropic and inotropic effects on the heart along with increased vascular resistance and a greatly diminished coronary flow reserve (CFR). These combined effects could intensify myocardial ischemia particularly in a person with moderate to severe atherosclerotic coronary artery disease (Mehta et al., 2004). Cocaine has been noted to induce a cellular level toxicity by reduction of mitochondrial respiration and increase mitochondrial generation of reactive oxygen species (ROS) in animals. Some reports reveal that production of ROS plays an important role in acute cocaine immunotoxicity and hepatotoxicity, both in vitro and in vivo (Aoki et al., 1997; Dıeź -Fernández et al., 1999; Pacifici et al., 2003). Cunha- Oliveira et al. (2013) and Yuan and Acosta (2000) reported that animals’ exposure to cocaine induced mitochondrial dysfunction in liver, brain and heart tissues to varying degrees. Chronic cocaine exposure followed by abrupt termination of cocaine increases the plasma corticosterone which leads to immune system failure and susceptible infections during the initial stage of withdrawal (Avila et al., 2004). Finally, cocaine has also been reported to cause death or seizure in male Sprague-Dawley rats due to primary respiratory arrest or direct cardiac toxicity (Tseng et al., 1993).

2.4.2.2 MAP and MDMA In general MAP and MDMA exert their effects by releasing catecholamine and serotonin through the transporters and blocking reuptake at presynaptic terminals (Rudnick and Wall, 1992; Sulzer et al., 1995; Mlinar et al., 2003). MAP mainly exerts its effects through the dopamine transporter (DAT) which enters the nerve terminals by passive diffusion and through the transporter at the nerve terminal (Zaczek et al., 1991). Conversely, MDMA has similar mechanisms mostly at the serotonin transporter (SERT) in some mammals only (Rudnick and Wall, 1992). However, MDMA can affect different species in different ways: in some mice models it mainly acts on the dopaminergic system (Easton and Marsden, 2006); in rats and non-human primates predominantly at serotonergic system (Rudnick and Wall, 1992). In certain situations, both MAP and MDMA act at the norepinephrine transporter (NET) and Vesicular monoamine transporter 2 (VMAT 2) (Jones et al., 1998). Animals exposed to MAP and MDAM may suffer several symptoms that range from being temporary to long- lasting. The most documented effects on animals were monoamine inhibition, neurotoxicity, apoptosis, oxidative stress, mitochondrial dysfunction, functional deficits and some behavioral

changes (Davidson et al., 2001; Lyles and Cadet, 2003; Krasnova and Cadet, 2009; Parrott, 2013; Karuppagounder et al., 2014).

2.4.2.2.1 Neurotoxicity MAP acts on both dopaminergic and the serotonergic systems and this results in an accumulation of DA in the synapse followed by long-term dopamine depletions (Krasnova and Cadet, 2009). MDMA has similar effects on the serotonin system (Rudnick and Wall, 1992). Bogen et al. (2003) have concluded that MDMA inhibits synaptosomal and vesicular uptake of both serotonin and dopamine. High dose chronic exposure of MAP to monkeys and rats caused toxicity to both DA and 5-HT terminals (Melega et al., 1997; Armstrong and Noguchi, 2004). Nerve terminal loss is evidenced by a loss of DAT, SERT, and VMAT 2 on the membrane surface (Guilarte et al., 2003). Both acute high doses and chronic smaller doses of administered MDMA can lead to a noticeable reduction in markers for serotonin across brain regions. These serotonergic changes could lead to definite neurotoxic changes in animals (Parrott, 2013). There is an evidence that amphetamine and MAP can produce serious hyperthermia followed by neurotoxicity in rats during the waking cycle, and the serum threshold level of the compounds required to produce the outcomes were similar to humans (Levi et al., 2012). In mice the hyperthermia and neurotoxic effects of MDMA were greater in older animals and subsequent dopaminergic damage compared to younger animals (Reveron et al., 2005).

2.4.2.2.2 Apoptosis and oxidative stress Initially it was assumed that amphetamines were only toxic to nerve terminals, but some recent studies have consistently reported that apoptotic cells in mouse brains emerge following administration of amphetamines (Jayanthi et al., 2004; Krasnova et al., 2005). Amphetamines’ brain neurotoxic effects occurred due to the increased production of reactive oxygen and nitrogen species (ROS and RNS, respectively), and LPO (Fleckenstein et al., 2007; Yamamoto et al., 2010). Amphetamines can cause an up-regulation of pro-apoptotic genes and down- regulation of anti-apoptotic genes due to the generation of ROS, and direct apoptosis activity (Thiriet et al., 2001; Deng et al., 2000). Some findings reveal that oxidative stress plays a role in MDMA-induced neurotoxicity where MDMA administration followed by increase of the LPO, an indicator of free radical damage (Gonçalves et al., 2014). Chronic exposure to MDMA results in a long-term depletion of brain 5-HT in rodents and non-human primates, and as a consequence increased free radical formation (oxidative stress) (Gudelsky and Yamamoto, 2008). Some evidence emerged that both acute overdose and chronic low dose of MAP exposure results in apoptosis or cell death, and oxidative stress in the rat cortex

and striatum and mouse hippocampus (Acikgoz et al., 2000; Gluck et al., 2001; Gonçalves et al., 2014). These neuropathological effects induced by MAP mainly occur via the production of ROS and RNS which are the primary causes of neurotoxicity (Davidson et al., 2001).

2.4.2.2.3 Behavioral effects MAP and MDMA exposure to animals led to various behavioral changes, such as psychosis, impulsivity, impaired memory and aggression. These behavioral changes may be associated with the amphetamines’ nerve terminal damage, oxidative stress and apoptosis activity. In animals, both acute and chronic exposure to MAP produced impairments in cognitive function and object recognition memory (Bisagno et al., 2002; Kamei et al., 2006). Later developmental, juvenile and neonatal MAP exposure leads to permanent cognitive and behavioral deficits in adulthood (White et al., 2009, McFadden et al., 2012), impaired hippocampus-dependent spatial memory in rodents in the Morris water maze (Acevedo et al., 2006) and the Barnes maze (Williams et al., 2003). MDMA and MDMA-analogues was observed producing several behavioral changes in the rodents as a result of enhanced oxidative stress and mitochondrial abnormalities (Karuppagounder et al., 2014). In general, the behavioral effects of MAP in animal model species depend on several variables such as strain, age, pre-treatment and metabolism (Good and Radcliffe, 2011).

2.4.3 Toxic effects on humans 2.4.3.1 Cocaine Abuse of cocaine leads to several risky medical complications for people, apart from its potential addiction properties. The toxic effects of cocaine varied little according to dosage, route of administration, age, and gender (Gili et al., 2014; Kohtz et al., 2010; Walker et al., 2009). So far, several studies have well documented cocaine’s neurological impact on humans (Kunisaki and Augenstein, 1994). In recent years, many cases have reported the rise in cardiovascular complications, such as unstable angina pectoris, acute myocardial infarction (AMI), aortic dissection, and infectious endocarditis (Gili et al., 2014; Maraj et al., 2010; Pozner et al., 2005). Fetal exposure to cocaine causes substantial prematurity, placental abruption, and premature rupture of the membrane (PROM) (Addis et al., 2001). Generally, chronic abuse of cocaine creates changes in people due to changes in the mesocorticolimbic dopamine system innervating ventral striatum, medial and orbitalfrontal cortex, nigrostriatal dopamine system and sensor motor (Hanlon et al., 2010; Volkow et al., 2008). While aberrant activation cell death in the adult human brain remains inconclusive, the apoptotic effects in cultured cells and in the developing brain are well documented. Overall the aberrant cell death indicates cocaine involvement of greater oxidative stress (Álvaro- 22

Bartolomé et al., 2011), it is also evident that prenatal cocaine exposure could delay children’s speech and language development (Cone-Wesson, 2005). Chronic cocaine use linked to cognitive impairment by altering sustained attention, response inhibition, lack of memory function, reward-based decision-making and changes in psychomotor performance (Spronk et al., 2013). In cultured cells, cocaine decreased the production of brain-derived, neurotrophic factor (BDNF) and this could make neurons more vulnerable to cocaine’s toxicity. Consequently, cocaine-induced central nervous system damage may occur in chronic abusers and prenatal exposure to cocaine (Yan et al., 2007). Chronic cocaine abuse leads to an imbalance in metabolic fat intake and storage process (physical side effects), can trigger weight gain and then withdrawal symptoms during the recovery period (Ersche et al., 2013).

2.4.3.2 MAP and MDMA The mode of action and effects of MAP and MDMA on human beings are very similar to animals. However, there have been several studies on human beings, which reported MAP and MDMA as having specific outcomes. Kim et al. (2006) demonstrated that lasting depletions occur in prefrontal grey matter density using magnetic resonance imaging (MRI) and impairment in frontal executive function even after more than 6 months’ abstinence from MAP abuse. Another study using positron emission tomography (PET) indicated a loss of dopamine transporters (DAT) in the striatum, which was related to impairments in memory and motor function (Volkow et al., 2001). One study showed lasting depletions in striatal DAT following an average of 3 years of abstinence (McCann et al., 1998). This leads to the question of whether the loss of DAT symbolizes an adaptive down-regulation or nerve terminal damage, which may be long-term. The effects of MAP are not limited to DAT and in fact, one analysis exhibited significantly decreased serotonin transporter density in abstinent MAP abusers associated with elevated levels of aggression (Sekine et al., 2006). In the case of MDMA, the effects on people are mostly indicated by the decrease in serotonin reuptake transporter rather than DAT (Parrott, 2013; Buchert et al., 2003; McCann et al., 2005). Several studies have been published on MAP’s and MDMA’s behavioral effects on human beings. MDMA regular users exhibit impaired memory compared to non-users (Parrott and Lasky, 1998) and also when compared to other polydrug users who had never used MDMA (Morgan et al., 1998). Regular use has also been reported to result in depression, anxiety, psychosis, and impulsivity. However, in a prospective study, MDMA users displayed more psychiatric symptoms than other drug users or non-users, confirming that MDMA plays a direct role (Parrot, 2006). Chronic amphetamine and MDMA use can lead to deficits in memory, attention and decision-making abilities in humans. Long-term use can also result in what has been termed amphetamine-induced psychosis (Kokkinidis and Anisman, 1981),

naturally consisting of paranoia, hallucinations, behavioral disorder, muscle tension, involuntarily teeth clenching, blurred vision, rapid eye movement, faintness, sweating, increasing blood pressure and heart disease (King & Ellinwood, 1997; Mas et al., 1999; Reichel et al., 2014). Impairment of neuropsychological functions including frontal executive function, memory recall impairment, and attention are observed in human chronic MAP abusers (Volkow et al., 2001; Sim et al., 2002; Kim et al., 2006). Very often a high MDMA dosage may cause cognitive and psychiatric symptoms, and these symptoms do not lend themselves readily to reversal upon quitting drug use (Karlsen et al., 2008). In addition long-term exposure to MAP causes geno-toxicity which is linked to changing frequencies in the micronucleus and SCE (sister chromatid exchange) Yu et al. (1999). Yamamoto and Raudensky (2008) believe that MAP and MDMA induced oxidative stress, excitotoxicity and mitochondrial dysfunction appear to play a major role in human neurotoxicity. However, direct, conclusive evidence of cellular toxicity/structural damage of ATSs in humans is yet to emerge.

2.4.4 Toxic effects on plants Only a few studies have examined the effects of illicit drugs on plants. However, Macht and Livingston, (1922) studied the effects of cocaine and how it decreased the growth of young roots of Lupinus albus. They suggested that although cocaine is most toxic in terms of what it does to animal tissues, it is comparatively less toxic to root growth. McCalla and Haskins (1964) have observed a toxic effect of morphine solution at 500 mg kg-1 to ladino clover seedlings in a complete nutrient culture. In fact, it is well known that organic amines, aliphatic amines and amino acids which comprise certain illicit drugs can inhibit photosynthesis, nitrogen fixation of plants and algae’s physiological process (Rosi-Marshall et al., 2014). In addition, treated wastewater and bio-solids contain considerable amounts of illicit drugs that are being used for crop cultivation (Pal et al., 2013). We as yet do not understand their effects on plants and associated herbivorous, which constitutes a potential research topic in the future.

2.5 Conclusion Illicit drugs’ environmental contamination is primarily due to human consumption and illegal disposal of waste materials from clandestine laboratories (Pal et al., 2013). Significant amounts of consumed parent drugs and/or their active metabolites are discharged into domestic wastewater through urine or faeces (Zuccato and Castiglioni, 2009). Currently available data indicate that BE, EME, MDMA, MAP, amphetamines and morphine are the most abundant residues in wastewater effluent. In line with the concentration levels documented in 24

these effluents, all six compounds occurred more frequently in surface water than other compounds. Despite their environmental concentrations being low, compounds such as BE, EME, morphine, amphetamine, and MDMA have potent biological activities and their presence as complex mixtures in surface waters may be toxic to aquatic organisms and human beings (Pal et al., 2013; Zuccato and Castiglioni, 2009). The potential effects of these compounds on the environment are largely unknown and no guideline values are available for what the acceptable level of illicit drugs and their metabolites in surface water should be. Our review of literature indicates that illicit drugs and their metabolites are mostly reported in European countries, the USA, Canada and Australia, but there is lack of reports available on Asia, Africa and the Oceanic region (Pal et al., 2013; Griffiths et al., 2008). The environmental fate of illicit drugs particularly in water streams depends on various factors such as compounds’ properties, seasonal variations and techniques used for removing illicit drugs in wastewater treatment plants. Until now very little information is available on the fate of illicit drugs in solid matrices, and therefore it is difficult to make a conclusion. Finally, the toxic effects of illicit drugs demonstrate that aquatic biota and vertebrates (animals and human beings) are all sensitive to illicit drugs. However, the concentration necessary to result in serious ecological damage requires further investigation.

Table 2.2 Illicit drugs’ concentrations in environmental systems

Environmental matrices Cocainics, opioids and Amphetamines and ecstasy References cannabinoids (ng L-1) groups (ng L-1)

Wastewater Influent 18.8 – 7500 2.2 -15380 Pal et al., 2013; Bijlsma et al., 2009; Effluent 7.5 - 3425 1.0 -10955 Huerta-Fontela et al., 2008a

Surface water 0.1 - 316 0.4 - 309 Pal et al., 2013; Bueno et al., 2011; Postigo et al., 2010

Drinking water/tape water 0.1 - 7.4 Jurado et al., 2012; Boleda et al., 2011; Valcárcel et al., 2012; Pal et al., 2013

Sewage sludge 0.1 - 300 µg Kg-1 Kaleta et al., 2006; Jones-Lepp and Stevens, 2007; Mastroianni et al., 2013

Atmospheric air 2.49 - 2800 pg m-3 1.05 - 10 pg m-3 Cecinato et al., 2009a; Viana et al., 2011; Pal et al., 2013

Table 2.3 Aquatic toxicity data for illicit drugs

Compound Species Trophic Effects/ Test Methods/approaches References group end point concentrations/ duration

Cocaine Danio rerio Fish Alteration in 1- 20 mg L-1 Conditioned place Darland and dopaminergic preference (CPP) Dowling, 2001 signalling in the brain Cocaine Dreissena Inverti. Primary DNA 40 ng L-1, Single cell gel Binelli et al., 2012 polymorpha damage; 220 ng L-1 and 10 electrophoresis (SCGE) increase in micro µg L-1; assay, the apoptosis nucleated cells 96 hr frequency evaluation & the and apoptosis micronucleus assay (MN test) and neutral red retention assay (NRRA) -1 Cocaine Tetrahymena Mono Mitogenic effects 10 and 20 µg mL ; Computerized image Stefanidou et al., (salt & freebase) pyriformis cellular due to stimulation 1 or 2 hr analysis system (CIAS) 2002 organisms of mitotic process and DNA synthesis.

Cocaine Anguilla anguilla Fish Bioaccumulation 20 ng L-1 High-pressure liquid Capaldo et al., 2012 in tissues, chromatography (HPLC) possible risk to method fish health in particular reproduction -1 Benzoylecogonine D. polymorpha Inverti. Oxidative stress- 0.5 and 1.0 µg L ; SCGE assay, the DNA Parolini et al., 2013 (BE) increases lipid 14 days diffusion assay and MN test peroxidation for DNA injuries, and NRRA (LPO), protein assay for cytotoxicity carbonylation (PCC), and DNA damages. Decreases of lysosomal membrane stability and imbalances of defence enzyme activities -1 Benzoylecogonine D. polymorpha Inverti. Alteration in 0.5 and 1.0 µg L ; Proteomics analysis -2-DE Binelli et al., 2013 (BE) important 14 days and mass spectrometry metabolic protein

profiles in gill analysis (RP-UPLC ESI- cells LTQ-Orbitrap) -1 Ecgonine methyl ester D. polymorpha Inverti. Weakening of 0.15 and 0.5µg L ; Multi-biomarker approach, Parolini and Binelli, 2013 (EME) lysosome 14 days in-vivo membranes, inactivation of defence enzymes, increases in LPO, PCC and DNA fragmentation Amphetamine sulfate Oncorhynchus Fish and Cytotoxic effects 3 hr, 24 hr Multi- centre evaluation of in Lilius et al., 1994 mykiss and Inverti. vitro cytotoxicity (MEIC) Daphnia magna Morphine sulfate Pseudopleurone Fish Plasma morphine 6 hr, 100 hr Enzyme-linked Newby et al., 2006 ctes americanus concentration immunosorbent assay and variations (ELISA) O. mykiss between fish. Lower disposition of morphine in fish than mammals

Morphine Elliptio Inverti. Reduction of 0 – 28.5 mg L-1 Fluorescamine method Gagné et al., 2006 complanata phagocytosis activity in hemolymph 9 -1 Δ - D. polymorpha Inverti. Oxidative stress 0.05 and 0.5µg L ; In vivo multi-biomarker Parolini and tetrahydrocannabinol induced LPO, 14 days approach Binelli, 2014 (THC) PCC and DNA damage 9 -1 Δ - D. rerio Fish Hypolocomotor 30 and 50 mg L ; Novel tank test Stewart and tetrahydrocannabinol effects/ 20 min Kalueff, 2014 (THC) anxiogenic-like response Heroine D. rerio Fish Increase in 15 and 25 mg L-1, Novel tank test Stewart and locomotion 20 min Kalueff, 2014

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Chapter 3

Illicit drugs - emerging contaminants in an urban environment

Abstract This work investigates the distribution pattern and occurrence of methamphetamine (MAP), 3, 4-methylenedioxymethamphetamine (MDMA), pseudoephedrine (PSE), cocaine (COC) and metabolites BE and ecgonine methyl ester (EME) in wastewaters, sewage sludge, surface waters and sediments from specific regions of South Australia. The results of this study revealed the presence of 3 out of 6 drugs of abuse targeted in wastewaters at concentrations ranging from 12 to 1670 ng L-1. MAP was the only test compound detected in sewage sludge (2 µg kg-1 dry samples). In surface waters MAP, MDMA and BE were recorded in 4 out of 20 test locations with a concentration of 5 to 11 ng L-1. Hence, water from wastewater treatment plants (WWTPs) could be primary potential source of illicit compounds that contaminate the environment. Although the concentrations of these contaminants in the environment are small, their impact on aquatic organisms and risk to human health cannot be ignored.

Keywords: Contaminants; Wastewater; Illicit drugs; Amphetamine-type stimulants; Cocaine

3.1 Introduction In recent years, illicit drugs and their metabolites have been identified globally in various environmental systems, and have been reported as being the latest group of emerging contaminants (ECs). ECs are substances that are not currently recognized as an impairment- causing agent in water systems, but they have the following characteristics that impact on the reliability of water: bioaccumulation; toxic effects; and persistence in the environment (Boles and Wells, 2010). The primary source of these residues in wastewater, and eventually the environment, is human consumption, since the major portion of the drugs and their metabolites are excreted and enter the sewage system. A secondary way in which they appear is by accidental or deliberate disposal from clandestine drug laboratories (Kasprzyk-Hordern et al., 2010; Boleda et al., 2011; Pedrouzo et al., 2011). Conventional wastewater treatment processes only partially effective in the removal of drug and metabolite residues (Zuccato et al., 2008). As a consequence, treated wastewaters have been reported to carry detectable

amounts of drugs that can contaminate surface and drinking waters if they are used for irrigation or recycled (Valcárcel et al., 2012; Pal et al., 2013). Recently, a number of countries in Europe, the United States and Canada have studied the drugs of abuse and their metabolites in wastewaters and surface waters (Postigo et al., 2011; Boleda et al., 2011; Valcárcel et al., 2012), sewage sludge (Kaleta et al., 2006; Jones- Lepp and Stevens, 2007) and atmospheric air (Viana et al., 2011; Cecinato et al., 2011). The main illicit drugs detected were COC and its metabolite BE ( cocainics), opioids, cannabinoids, and the ATS and MDMA. The global occurrence of illicit drugs and their metabolites in the environment has recently been reviewed by our group (Pal et al., 2013). In Australia, Irvine et al. (2011) reported the presence of MAP (2- 4108 ng L-1), MDMA (5-706 ng L-1), and BE (12- 103 ng L-1) in South Australian wastewaters while Lai et al reported the median daily per capita loadings of these drugs at a municipal wastewater plant in south east Queensland to be 21.9 g, 6.7 g and 20.3 g, respectively. Internationally as well as in Australia, ATSs are the second most commonly abused illicit drugs after cannabis. They are easy to manufacture in clandestine laboratories with commonly available precursors and chemicals (EMCDDA, 2008; ACC, 2011-12; UNODC, 2012). Meanwhile, in 2011-12, the detection of ATSs (excluding MDMA) at the Australian border both in terms of number and weight has increased and is the highest reported in the last decade (ACC, 2011 -12). Illicit manufacture of ATSs has been reported in more than 70 countries and the global market for the synthetic ATSs drugs evolved quickly (UNODC, 2012). Unlike other non-polar contaminants, these drugs are not readily sorbed to the subsoil or completely removed by treatment processes, which increases their potential to subsequently enter surface and groundwater (Jones-Lepp et al., 2004). Furthermore, ATSs and cocainics may have potent pharmacological activities and their presence as complex mixtures in water may cause adverse effects on aquatic organisms and people’s health (Pomati et al., 2006; Binelli et al., 2012). A steady increase in drug abuse in Australian communities (ACC, 2011-12) coupled with growing domestic usage of treated water (Mainali et al., 2011) may cause serious environmental problems in the future. To date, no systematic study has been published on the presence of illicit drugs in treated wastewater, sewage sludge, surface waters and sediments in South Australia. The aim of this study was to investigate the occurrence and distribution pattern of ATSs and cocainics in different environmental systems in specific regions of South Australia.

3.2 Materials and methods

3.2.1 Reagents and materials The target compounds MAP, MDMA, pseudoephedrine (PSE), cocaine (COC) and metabolites BE and ecgonine methyl ester (EME) were obtained from Forensic Science SA, Adelaide. Analytical grade methanol, ammonium acetate, acetic acid, chloroform, and acetonitrile were obtained from Sigma-Aldrich (Sydney, Australia). Water was purified using Elga-Purelab classic system (Thermo Fisher, Vic., Australia). Oasis MCX cartridges (150 mg/6 cc) for extraction were purchased from Waters Corporation (Milford, MA, USA) and Alltech Solid- Phase Extraction (SPE) vacuum manifold with 12 port system was used for the loading of water samples and drying of the cartridges.

3.2.2 Sample collection and analysis Sewage water (influent and effluent) and recovered activated sludge samples were collected from three metropolitan WWTPs, which together service the entire Adelaide metropolitan population of some 1.29 millions. Surface waters and sediments were collected from different rivers, lakes, and creeks in northern Adelaide (Fig. 3.1). All test samples were collected once between June 2012 and June 2013 (Table 3.2), as grab samples in amber coloured glass bottles. The pH of the wastewaters and surface waters was adjusted to 2 by adding 37% hydrochloric acid (HCl) to prevent degradation of the compounds, and then they were filtered through glass-fibre filters (Whatman, GF/C). Samples were stored for a maximum of 3 days in the dark at 4 °C before extraction and the final extracts were stored at -20 °C for 7 days maximum before analysis.

3.2.3 Analysis of illicit drugs and quality control The samples were analysed for selected illicit drugs and their metabolites with a previously documented method based on SPE and HPLC-MS (Castiglioni et al., 2006). All individual samples were extracted once and each extract was analysed twice by HPLC-MS. In brief, water samples (500 mL of wastewater, 1 L of surface water) were passed through conditioned SPE cartridge (Oasis MCX) followed by two consecutive elutions using methanol and 2% ammoniacal methanol. The eluate was evaporated under a stream of nitrogen at room temperature and the dried residue was redissolved in 200 µL methanol for analysis. The sediment and sludge samples were extracted following the method described by Pal et al. (2011).

The test compounds were separated by HPLC (Agilent 1100 series) equipped with an auto-sampler, binary pump system, a mass selective detector (Agilent1100), and chemstation software for data integration. ZORBAX Eclipse XDB-C18 150x4.6 mm, 5 µm column at 25 °C was employed. The mobile phase was composed of solvent A (20% methanol + 0.1% acetic acid + 10 mM ammonium acetate) and solvent B (90% methanol + 0.1% acetic acid + 10 mM ammonium acetate). The flow-rate was 0.8 mL min-1 and the injection volume was set at 10 µL. Positive polarity and electrospray ionization (ESI) were used. The scan was conducted in the mass range of 100 to 350 Daltons, EMV Gain was set at 3.0 fragmentor voltage was 120V, threshold was 0.0, and step size was 0.10. The spray chamber pressure for the nebulizer was 35 psig and drying gas 12.0 L min-1 (Pal et al., 2011). Regular tests were carried out to check the sensitivity, repeatability, chromatographic separation and calibration of the instrument. The analytical instrument’s performance was monitored by analysing the set of LC-grade water spiked with the standard mixture samples at 50 ng L-1 included between every five samples in the sample batch. The standards for calibration was prepared by serial dilution of the mixed working solution using deionised water resulting in individual concentrations ranging from 10 to 1000 ng L-1 for wastewater and 0.2 to 10 ng L-1 for surface waters. Fresh calibration standards in water were used for every new series of samples. The limits of quantitation (LOQ) were set 5 ng L-1 PSE, MAP and MDMA, 6 ng L-1 (COC and BE) and 10 ng L-1 for EME, after established the concentration of the analytes in distilled water that gives rise to peak height with an S/N of 10. The two most responsive transitions per compound were used for quantitation of analytes. In this extraction method the analytes average recoveries were 65 ± 2.3 (PSE) to 89 ± 1.6 (MAP) when n = 3. Method repeatability, evaluated as the relative standard deviation (RSD) of the replicate analysis of spiked (50 ng L-1) water samples was also satisfactory, with RSD values between 1.5 to 5.4%.

3.3 Results and discussion The results of the present study showed the presence of MAP, MDMA, and BE in the test samples. Table 3.1 summarises the distribution pattern and occurrence levels of illicit drugs and metabolites in wastewaters, surface waters, and sludge throughout the study area. The target compounds were found to be present in 3 of the Adelaide metropolitan area wastewater influents, in 2 effluents and in 1 sewage sludge sample. In surface waters and sediments collected from the areas surrounding the city of Adelaide, the drugs were identified in 4 locations out of 32. MAP, MDMA and BE were detected in all sewage influent waters. COC, EME and PSE were not detected in any of the samples. The concentration levels of target compounds in

influent water samples ranged between 12 and 1670 ng L-1 (Table 3.1). The highest residues of MAP and MDMA were observed in WWTP#1 and BE in WWTP#3. MAP and MDMA concentrations recorded in this study are close to those previously reported in Australia (Irvine et al., 2011), and those documented in Spain and the US (Huerta-Fontela et al., 2007; Chiaia et al., 2008; Bijlsma et al., 2009; Bueno et al., 2011). However, the BE recorded was much less than Italy, Switzerland, Germany, Belgium, Canada and the UK (Castiglioni et al., 2006; Hummel et al., 2006; Kasprzyk-Hordern et al., 2009; Berset et al., 2010; Metcalfe et al., 2010; van Nuijs et al., 2011). These results indicate that ATS, MDMA and to a lesser extent COC are commonly abused illicit drugs in Adelaide areas. In effluent waters, 4 of the 6 target compounds were not detected in WWTP#1 and WWTP#2, and only MAP was observed in WWT#3. The most abundant compound was MAP, present in 100% of the treated waters, followed by MDMA in 67% of the samples, with concentrations reaching 135 ng L-1 and 68 ng L-1, respectively (Table 3.1). The presence of these compounds in effluent waters shows that the contaminants in the WWTPs were not completely removed. BE was completely eliminated from wastewater, which may be due to a higher removal efficiency or lower input load of compounds, whereas 93% of MAP and 75% of MDMA were removed. This is consistent with previous studies (Huerta-Fontela et al., 2008a; Kasprzyk-Hordern et al., 2009; Boles and Wells, 2010; Postigo et al., 2010), which indicated that the contaminant removal efficiency of WWTP was influenced by several factors, such as the compound’s physicochemical properties, source water quality parameters, treatment methods used, and biological population in water (Huerta-Fontela et al., 2008b; Postigo et al., 2010). With reference to sewage sludge, only MAP was detected and only in WWTP#1 with the concentration being up to 2 µg kg-1 in dry samples (Table 3.1) which is of interest to note that this plant received the highest inflow of MAP (1670 ng L-1) of the three plants studied.,. This recorded concentration is lower than earlier studies, where: firstly, amphetamine (5 - 300 µg kg-1 dry samples) was found in Austrian sewage sludge by Kaleta et al. (2006); and secondly, MAP (4 μg kg-1 dry weight) was found in bio-solids from Los Angeles, by Jones-Lepp and Stevens (2007). The chemical and physical composition of sewage sludge can vary on a daily basis, even within a treatment plant, and this poses an analytical challenge for determining contaminants (Jones-Lepp and Stevens, 2007). MAP has been shown to be stable in nature (Janusz et al., 2003; Pal et al., 2011), therefore it is perhaps not surprising to find that it can survive water treatment processes and persist in sludge. Three of the target illicit substances (MAP, MDMA and BE) were found in surface waters collected from 4 locations out of the 20 tested at concentrations ranging from 5 to 11 ng L-1; with MDMA detected in all 4 test sites followed by MAP and BE in 1 site each. The observed concentration of MDMA in surface water is comparable to those reported earlier in

similar monitoring studies (Huerta-Fontela et al., 2008b; Zuccato et al., 2008; Valcárcel et al., 2012). However, the recorded concentration of BE is much less compared to European countries, such as Italy (Zuccato et al., 2008), Spain (González-Mariño et al., 2010; Bueno et al., 2011), Belgium (Gheorghe et al., 2008; van Nuijs et al., 2009) and the UK (Kasprzyk- Hordern et al., 2008, 2009) however, as we have shown above, the rate of usage of cocaine in Adelaide appears to be lower as well. The four test sites receive treated water from WWTP#1, consequently it is reasonable to surmise that wastewater contributes illicit compounds that contaminate surface waters. This supports the view illicit drugs are a group of potential environmental contaminants. Although the environmental concentrations of the detected illicit drugs and metabolites are small, due to their pharmacological activities and chronic exposure as mixtures of compounds in surface waters, they are potentially toxic to aquatic organisms and therefore may pose risk to human health (Zuccato and Castiglioni, 2009).

3.4 Conclusion The distribution pattern and occurrence of ATS and BE in wastewaters and surface waters in specific regions of Adelaide has been demonstrated. The analysis of influent, effluent and sewage sludge of three metropolitan WWTPs and surface waters collected in the adjoining Adelaide areas indicated that MAP, MDMA and BE were the most abundant and ubiquitous compounds. In general, our findings agrees with those of Irvine et al. (2011) and the World Drug Report (2011), which showed higher use of MAP and MDMA in Australia in comparison to other countries being monitored. Our research results also suggest that illicit compounds contaminating the environment could originate from WWTP treated water. The levels observed in surface waters and sewage sludge are sufficient to affect aquatic organisms and soil microbes in some respects.

Table 3.1 Concentration of illicit drugs and their metabolites in wastewaters, sewage sludge and surface waters in Adelaide surrounding areas.

Location / ID Sample Concentration of drug residues (ng L-1) (mean ± SD) Type Waste MAP MDMA PSE COC BE EME water WWTP#1 Influent 1670 ± 56 236 ± 15 nd nd 31 ± 7 nd Effluent 135 ± 7 68 ± 2 nd nd nd nd Sludge 2 ± 1 µg kg-1 nd nd nd nd nd WWTP#2 Influent 480 ± 71 112 ± 71 nd nd 12 ± 3 nd Effluent 23 ± 4 24 ± 2 nd nd nd nd Sludge nd nd nd nd nd nd WWTP#3 Influent 195 ± 21 68 ± 44 nd nd 47 ± 4 nd Effluent 10 ± 2 nd nd nd nd nd Sludge nd nd nd nd nd nd Mawson Lakes - Surface Panorama#1SW Water nd 9 ± 2 nd nd nd nd Mawson Lakes - Surface Shear water#3SW Water nd 5 ± 1 nd nd nd nd Mawson Lakes - Surface Cascades drive Water #12SW nd 8 ± 2 nd nd nd nd Pooraka Surface Lake#16SW Water 11 ± 2 8 ± 1 nd nd 7 ± 1 nd nd: not detected

(a)

(b)

(c)

Figure 3.1. Geographical locations of the sampling points

(a)

(b)

(c)

3.5 References ACC (Australian Crime Commission), Illicit drug data report 2011-2012. http://www.crimecommission.gov.au/publications/illicit-drug-data-report/illicit-drug- data-report-2011-12 [Accessed on June 15 2013]. Berset JD, Brenneisen R, Mathieu C. Analysis of licit and illicit drugs in waste, surface and lake water samples using large volume direct injection high performance liquid chromatography-electrospray tandem mass spectrometry (HPLC-S/MS). Chemosphere 2010; 81:859-66. Bijlsma L, Sancho JV, Pitarch E, Ibáñez M, Hernádez F. Simultaneous ultra high pressure liquid chromatography-mass spectrometry determination of amphetamine like stimulants, cocaine and its metabolites, and a cannabis metabolite in surface water and urban wastewater. J Chromatogr A 2009; 1216:3078-89. Binelli A, Pedriali A, Riva C, Parolini M. Illicit drugs as new environmental pollutants: cyto- genotoxic effects of cocaine on the biological model Dreissena polymorpha. Chemosphere 2012; 86:906 -11. Boleda MR, Huerta-Fontela M, Galceran MT, Ventura F. Evaluation of the presence of drugs of abuse in tap waters. Chemosphere 2011; 84:1601-7. Boles TH, Wells MJM. Analysis of amphetamine and methamphetamine as emerging pollutants in wastewater and wastewater-impacted streams. J Chromatogr A 2010; 1217:2561-8. Bueno MJM, Uclés S, Hernando MD, Fernández-Alba AR. Development of a solvent free method for the simultaneous identification/quantification of drugs of abuse and their metabolites in environmental water by LC-MS/MS. Talanta 2011; 85:157-66. Castiglioni S, Zuccato E, Crisci E, Chiabrando C, Fanelli R, Bagnati R. Identification and measurement of illicit drugs and their metabolites in urban wastewater by liquid chromatography-tandem mass spectrometry. Anal Chem 2006; 78:8421-9. Cecinato A, Balducci C, Guerriero E, Sprovieri F, Cofone F. Possible social relevance off illicit psychotropic substances present in the atmosphere. Sci Total Environ 2011;412– 413:87–92. Chiaia AC, Banta-Green C, Field J. Eliminating solid phase extraction with large volume injection LC/MS/MS: analysis of illicit and legal drugs and human urine indicators in US wastewaters. Environ Sci Technol 2008; 42:8841-8. EMCDDA (European Monitoring Centre for Drugs and Drug Addiction). Annual report: the state of the drugs problem in Europe. http://www.emcdda.europa.eu/html.cfm/index64151EN.html2008 . [Accessed on May 25, 2013].

Gheorghe A, van Nuijs A, Pecceu B, Bervoets L, Jorens PG, Blust R. Analysis of cocaine and its principal metabolites in waste and surface water using solid-phase extraction and liquid chromatography-ion trap tandem mass spectrometry. Anal Bioanal Chem 2008; 391:1309-19. González-Mariño I, Quintana JB, Rodríguez I, Cela R. Determination of drugs of abuse in water by solid phase extraction, derivatisation and gas chromatography ion trap tandem mass spectrometry. J Chromatogr A 2010; 1217:1748-60. Huerta-Fontela M, Galceran MT, Martin-Alonso J, Ventura F. Occurrence of psychoactive stimulatory drugs in wastewaters in north-eastern Spain. Sci Total Environ 2008a; 397: 31-40. Huerta-Fontela M, Galceran MT, Ventura F. Stimulatory drugs of abuse in surface waters and their removal in a conventional drinking water treatment plant. Environ Sci Technol 2008b; 42:6809-16. Huerta-Fontela M, Galceran MT, Ventura F. Ultraperformance liquid chromatography tandem mass spectrometry analysis of stimulatory drugs of abuse in wastewater and surface water. Anal Chem 2007; 79:3821-9. Hummel D, Löffler D, Fink G, Ternes TA. Simultaneous determination of psychoactive drugs and their metabolites in aqueous matrices by liquid chromatography mass spectrometry. Environ Sci Technol 2006; 40:7321-8. Irvine RJ, Kostakis C, Felgate PD, Jaehne EJ, Chen C & White JM. Population drug use in Australia: A wastewater analysis. Forensic Sci Int 2011; 210:69-73. Janusz A, Kirkbride KP, Scott TL, Naidu R, Perkins MV, Megharaj M. Microbial degradation of illicit drugs, their precursors, and manufacturing by-products: implications for clandestine drug laboratory investigation and environmental assessment. Forensic Sci Int 2003; 134:62-71. Jones-Lepp TL, Alvarez DA, Petty JD, Huckins JN. Polar organic chemical integrative sampling and liquid chromatography-electrospray/ion trap mass spectrometry for assessing selected prescription and illicit drugs in treated sewage effluents. Arch Environ Contam Toxicol 2004; 47:427-39. Jones-Lepp TL, Stevens R. Pharmaceuticals and personal care products in biosolides/sewage sludge: the interface between analytical chemistry and regulation. Anal Bioanal Chem 2007; 387:1173-83. Kaleta A, Ferdig M, Buchberger W. Semiquantitative determination of residues of amphetamine in sewage sludge samples. J Sep Sci 2006; 29:1662- 6. Kasprzyk-Hordern B, Dinsdale RM, Guwy AJ. Multiresidue methods for the analysis of pharmaceuticals, personal care products and illicit drugs in surface water and

wastewater by solid-phase extraction and ultra-performance liquid chromatography- electrospray tandem mass spectrometry. Anal Bioanal Chem 2008; 391: 1293-308 Kasprzyk-Hordern B, Dinsdale RM, Guwy AJ. The removal of pharmaceuticals, personal care products, endocrine disruptors and illicit drugs during wastewater treatment and its impact on the quality of receiving waters. Water Res 2009; 43:363-80. Kasprzyk-Hordern B, Kondakal VVR, Baker DR. Enantiomeric analysis of drugs of abuse by chiral liquid chromatography coupled with tandem mass spectrometry. J Chromatogr A 2010; 1217:4575-86. Mainali B, Ngo HH, Guo W, Pham TTN, Johnston A. Feasibility assessment of recycled water use for washing machines in Australia through SWOT analysis. Resources, Conservation and Recycling 2011; 56: 87-91. Metcalfe C, Tindale K, Li H, Rodayan A, Yargeau V. Illicit drugs in Canadian municipal wastewater and estimates of community drug use. Environ Pollut 2010; 158:3179-85. Pal R, Megharaj M, Kirkbride KP, Heinrich T & Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor, and by-products in soil. Chemosphere 2011; 85:1002-1009. Pal R,Megharaj M,Kirkbride KP,Heinrich T & Naidu R. Illicit drugs and the environment-A review. Sci Total Environ 2013; 463-464:1079-1092. Pedrouzo M, Borrull F, Pocurull E, Marcé RM. Drugs of abuse and their metabolites in waste and surface waters by liquid chromatography tandem mass spectrometry. J Sep Sci 2011; 34:1091-101. Pomati F, Castiglioni S, Zuccato E, Fanelli R, Rossetti C, Calamari D. Effects of environmental contamination by therapeutic drugs on human embryonic cells. Environ Sci Technol 2006; 40:2442-7. Postigo C, de Alda MJL, Barceló D. Drugs of abuse and their metabolites in the Ebro river basin: occurrence in sewage and surface water, sewage treatment plants removal efficiency, and collective drug usage estimation. Environ Int 2010; 36:75-84. Postigo C, de Alda MJL, Barceló D. Evaluation of drugs of abuse and trends in a prison through wastewater analysis. Environ Int 2011; 37:49-55. United Nations Office on Drugs and Crime (UNODC) 2012, Global Smart Update 2012, vol. 7, UNODC, Vienna. Valcárcel Y, Martínez F, González-Alonso S, Segura Y, Catalá M, Molina R. Drugs of abuse in surface and tap waters of the Tagus river basin: heterogeneous photo-Fenton process is effective in their degradation. Environmental International 2012; 41:35-43. van Nuijs ALN, Mougel JF, Tarcomnicu I, Bervoets L, Blust R, Jorens PG, et al. A one year investigation of the occurrence of illicit drugs in wastewater from Brussels, Belgium. J Environ Monit 2011; 13:1008-16.

van Nuijs ALN, Pecceu B, Theunis L, Dubois N, Charlie C, Jorens PG. Can cocaine use be evaluated through analysis of wastewater? A nation-wide approach conducted in Belgium. Addiction 2009; 104:734-41. Viana M, Postigo C, Querol X, Alastuey A, de Alda MJL, Barceló D. Cocaine and other illicit drugs in airborne particulates in urban environments: a reflection of social conduct and population size. Environ Pollut 2011; 159:1241-7. Zuccato E, Castiglioni S, Bagnati R, Chiabrando C, Grassi P, Fanelli R. Illicit drugs, a novel group of environmental contaminants. Water Res 2008; 42:961-8 Zuccato E, Castiglioni S. Illicit drugs in the environment. Philos Trans R Soc A 2009; 367: 3965-78

Table 3.2 (supplementary) : Sample collection details

Sample Sample Location pH Weather Sample ID collection Extraction date Max. Relative Rainfall Date Temperature humidity (%) (mm) °C WWTP#1 14/06/13 Influent 6.7 15 70 0 14/06/13 Effluent 7.1 Sludge 6.4 WWTP#2 29/05/13 Influent 6.6 16 72 0 30/05/13 Effluent 7.0 Sludge 6.5 WWTP#3 23/05/13 Influent 6.8 14 71 0 23/05/13 Effluent 7.2 Sludge 6.5 1SW 11/07/12 Panorama, 7.3 15 71 0 12/07/12 Mawson Lakes 2S 11/07/12 Panorama, 7.2 15 71 0 12/07/12 Mawson Lakes

3SW 9/06/12 Shearwater, 7.3 12 51 0 11/06/12

Mawson Lakes 4S 9/06/12 Shearwater, 7.4 12 51 0 11/06/12 Mawson Lakes 5S 9/06/12 Hamilton place, 7.5 13 52 0 11/06/12 Mawson Lakes 6SW 9/06/12 Hamilton place, 7.1 13 52 0 11/06/12 Mawson Lakes 7SW 9/06/12 Dry creek, 7.1 14 52 0 11/06/12 Main north Road 8S 9/06/12 Dry creek, 6.8 14 52 0 11/06/12 Main north Road 9SW 27/06/12 Shoalhaven lake, 7.1 15 70 0 28/06/12 Mawson Lakes 10SW 18/06/12 Greenfield, 6.9 13 79 0 20/06/12 Salibury Hwy 11S 18/06/12 Greenfield, 7.5 13 79 0 20/06/12 Salibury Hwy 12SW 18/06/12 Cascades Drive, 7.0 12 76 0 20/06/12 Mawson Lakes 13S 18/06/12 Cascades Drive, 7.4 12 76 0 20/06/12 Mawson Lakes 14S 18/06/12 Isala circuit, 7.1 12 74 0 20/06/12

Mawson Lakes 15SW 18/06/12 Isala circuit, 7.4 12 74 0 20/06/12 Mawson Lakes 16SW 19/06/12 Main north road, 6.5 15 75 0 21/06/12 Pooraka 17S 19/06/12 Main north road, 5.6 15 75 0 21/6/12 Pooraka 18S 30/10/12 Wynn vale Dam, 7.1 24 53 0 02/11/12 Wynn Vale 19SW 30/10/12 Wynn vale Dam, 6.6 24 53 0 02/11/12 Wynn Vale 20SW 30/10/12 Ross Rd, 6.7 23 53 0 02/11/12 Golden Grove 21S 30/10/12 Ross Rd, 7.2 23 53 0 02/11/12 Golden Grove 22SW 30/10/12 Little Para River, 6.7 22 56 0 02/11/12 Carobroke Park 23SW 15/11/12 Little Para Reservoir 6.6 29 48 0 17/11/12 24S 15/11/12 Little Para Reservoir 7.2 29 48 0 17/11/12 25SW 15/11/12 Little Para River ,Little 6.6 28 48 0 17/11/12 Para Linear Park

26SW 15/11/12 Little Para River ,Happy 6.8 28 48 0 17/11/12 Home Reserve 27SW 15/11/12 Muffin Reserve, 6.8 29 45 0 17/11/12 Elizabeth 28SW 11/07/12 Lakeside Circuit, 7.1 15 71 0 12/07/12 Northgate 29SW 25/11/12 River Torrens, 6.7 30 43 0 26/11/12 Adelaide 30S 25/11/12 River Torrens, 7.2 30 43 0 26/11/12 Adelaide 31SW 25/11/12 River Torrens, 6.8 29 44 0 26/11/12 Port Rd 32SW 15/09/12 Morialta Falls, 7.0 20 56 0 17/09/12 Morialta Conservation Park

Chapter 4

Sorption and desorption patterns of amphetamine-type substances (ATS) in different soils – the influence of soil properties

Abstract

This paper examines the sorption kinetics, isotherms, sorption and desorption patterns of methamphetamine (MAP) and 3, 4- methylenedioxymethamphetamine (MDMA) in three different soils using batch equilibration experiments. Both the target compounds reached equilibrium within 12 h. The sorption data were analysed employing the Langmuir and Freundlich models, and the results showed that the Freundlich model best fit and described the sorption process of MAP and MDMA in three test soils. The order of sorption in treatments was: MAP ˃ MAP in mixture ˃ MDMA

in mixture ˃ MDMA. The sorption coefficient (Kd) was positively correlated with soils’ organic carbon (OC), dissolved organic carbon (DOC) and clay for MAP, while for MDMA it was clay, OC and DOC. Desorption showed following order: MAP ˃ MAP in mixture ˃ MDMA ˃ MDMA in mixture. These findings could give an insight into the sorption and desorption patterns and act as an information tool for the transport, fate and risk assessment of MAP and MDMA in the environment.

Keywords: Methamphetamine; MDMA; Soil; Sorption; Kinetics

4.1 Introduction

In the amphetamine-type substances (ATS) group, methamphetamine (MAP) and 3, 4-methylenedioxymethamphetamine (MDMA) are popular drugs of abuse. The use of MAP and MDMA is currently demanding most of our attention regarding the synthetic illicit drugs, due to their adverse impact on people’s health and social welfare (Pal et al., 2011). Globally, ATS manufacture is common throughout East and South-East Asia, North America, Oceania and Africa (UNODC, 2008. In Australia, the ATS market is second only to cannabis and this trend may continue to grow (ACC, 2011-12). MAP continues to be the most widely manufactured ATS (68%) followed by MDMA. Mostly these drugs are manufactured domestically in clandestine laboratories using locally available chemicals through a variety of manufacturing processes (UNODC, 2008). 65

ATS occurrence in the natural environment has been reported in several regions of the world. The main pathways of ATS into the environment are excretion following human consumption and illegal disposal of clandestine waste materials into waste management facilities (Pal et al., 2013). Further, MAP and MDMA once released into the environment may undergo various soil processes such as degradation, sorption, leaching, and runoff (Pal et al., 2013). The presence of ATS compounds in sewage sludge, surface and ground water not only indicates these compounds’ environmental behaviour, but also warns us about the associated risk to people and wildlife. The occurrence of these types of relatively persistent and pharmacologically active pollutants in different environmental matrices and their ability to migrate to the groundwater system after accidental spill or as a consequence of waste disposal, is an important environmental concern which requires knowledge on sorption patterns under the influence of different soil characteristics. In addition, the sorption process of an organic contaminant is determined by their availability to be degraded by microorganisms, bioactivity, phytotoxicity, and biotransformation (Delle Site, 2001). Considering the environmental behaviour of ATS and current abuse level in Australia, the present study was aimed to investigate the sorption and desorption potential of MAP and MDMA in various Australian soils in terms of single and compound mixtures

4.2 Materials and methods

4.2.1 Chemicals Refer chapter 3.2.1

4.2.2 Soils The soil samples used in this study were collected from three different locations in South Australia: Mount Lofty (ML); Crafers West (CW), and Wallaroo (W). Surface soils (0 -15 cm) were collected separately in each location in polypropylene buckets and brought to the laboratory. The soils were air-dried at room temperature and screened to remove any plant parts or other debris. The well-homogenised and sieved soils (2 mm) were stored at 4 ºC for experimental use. The physico-chemical properties of the soils were analysed using standard analytical methods and the results are provided in Table 4.1. The test soils varied widely in terms of pH, electrical conductivity (EC), organic carbon (OC), clay content, and soil texture. The pH of the soils ranged between 5.3 (acidic), 7.2 (neutral), and 8.8 (alkaline) for ML, CW and W soil, respectively. The EC and cation exchange capacity (CEC) of the soils ranged between 30.3 – 810 µScm-1 and 1.3 – 4.9 meq-100g-1. The OC content in ML, CW, and W

soils were 3.1, 3.8, and 1.9%, respectively. These soils contained a high level of sand (66.8 – 82.7%), moderate level of clay (10.6 – 23.4%) and low level of silt (6.7 – 13.3%). Loamy sand and sandy loam texture was observed in the ML and CW soils while sandy clay loam texture was evident in the W soil.

4.2.3 Sorption and desorption test MAP and MDMA sorption and desorption patterns in three different South Australian soils were studied in accordance with Pal et al. (2012) tested method. This experiment was conducted in dark at room temperature using 50 mL screw cap tubes to avoid photo- -1 degradation. MAP and MDMA stock solutions (60 mg L ) were prepared in 0.01 M CaCl2. To investigate the equilibration time, the first stage of the sorption experiment was -1 conducted using two concentrations (10 and 60 mg L ) of MAP and MDMA in 0.01M CaCl2, over a range of equilibration times (0, 1, 2, 4, 6, 12, and 24 h). The second stage of the experiment was conducted at 12 h equilibration time and 5, 10, 20, 40 and 60 mg L-1 concentrations were used based on the first stage results. Five grams of soils were weighed into a 50 mL tube and the required amount of stock solution was added to achieve the desired concentration. Blanks without soils indicated no significant sorption loss on the tube walls. Additionally, control soils with a blank 0.01 M CaCl2 solution were maintained. All the experiments were conducted in duplicate. The vials were shaken in an electrical shaker to achieve equilibration and the supernatants were centrifuged at 3000 g for 15 min. The aliquots were passed through a 0.22 μm filter for direct analysis in HPLC-MS. Desorption experiments were conducted immediately after the sorption test, testing one concentration (60 mgL-1) only. The aliquots from the sorption test were discarded carefully without losing any soil particles and 20 mL of 0.01 M CaCl2 background solution were added. The vials were shaken for 12 h under the above experimental conditions following the similar sorption experimental procedures.

Table 4.1 Basic physico-chemical properties of the experimental soils

Soil pH Electrical Cation Organic Dissolved Particle size Textural B.E.T Clay minerals conductivity exchange carbon organic distribution class Surface (μScm-1) capacity (%) carbon Sand Silt Clay area (meq 100g-1) (µg mL-1) (%) (%) (%) (m2g-1) Mount Lofty 5.3 30.3 1.6 3.1 13.5 82.7 6.65 10.6 Loamy sand 1.5 Quartz low ˃ (ML) Albite ˃ Zeolite Rho ˃ Kaolinite

Crafers West 7.2 155 1.3 3.8 14.7 66.8 13.29 19.8 Sandy loam 12.8 Quartz ˃ (CW) Kaolinite ˃ Muscovite ˃ Zeolite X

Wallaroo 8.8 810 4.9 1.9 11.5 68.6 8.04 23.4 Sandy clay 19.2 Quartz low ˃ (W) loam Calcite ˃ Augite

4.2.4 Compounds extraction and analysis MAP and MDMA were extracted and analysed by HPLC-MS following the procedure described by Pal et al. (2012). Test compounds were extracted from soil using 40 mL of extraction solvents (chloroform: acetonitrile: methanol: acetic acid - 80:10:9:1) in two steps, 20 mL for each step. The supernatant was centrifuged and filtered using 0.22 µm Teflon filters. The combined 40 ml of the supernatants was dried under a nitrogen stream and reconstituted with methanol for analysis. The test compounds were determined by HPLC (Agilent 1100 series) equipped with an auto-sampler, binary pump system, a mass selective detector (Agilent1100), and Chemstation software for data integration: ZORBAX Eclipse XDB-C18 150x4.6 mm, 5µm columns at 25°C. The mobile phase contained two groups, solvent A (20% methanol + 0.1% acetic acid + 10 mM ammonium acetate) and solvent B (90% methanol + 0.1% acetic acid + 10 mM ammonium acetate). The flow-rate was 0.8 mL min-1and 10 µL of samples were injected following 26 min mobile phase total runtime. Positive polarity and electrospray ionization (ESI) techniques were employed in the detector. The scan was conducted in the mass range of 100 to 350, Gain EMV 3.0 Fragmentor 120, Threshold 0.0, and step size 0.10. The spray chamber pressure for the nebulizer was 35 psi and drying gas 12.0 L min- 1 (Pal et al., 2011).

4.2.5 Calculation of sorption parameters The sorption behavior of MAP and MDMA in three different soils was calculated in this experiment by using sorption isotherms and sorption coefficient. The sorption isotherms were obtained by plotting the amount of chemical sorbed per unit weight of soil (X) versus the amount of chemical per volume of solution at equilibrium (C). In this study Langmuir and Freundlich isotherm models were used for calculating the MAP and MDMA sorption isotherms. The applied Langmuir model equation was as follows:

= 1 + 𝑋𝑋𝑚𝑚 𝐾𝐾𝐾𝐾 𝑋𝑋 Where: 𝐾𝐾𝐾𝐾 K = Langmuir equilibrium constant C = Sorbent concentration in solution X = Amount adsorbed Xm = Max. amount adsorbed as C increases Data were fitted to the Langmuir linear regression equation: 1 = +

𝐶𝐶 𝐶𝐶 69 𝑋𝑋 𝑋𝑋𝑚𝑚 𝐾𝐾 𝑋𝑋𝑚𝑚

A plot of C/X vs. C yields a slope = 1/Xm and an intercept = 1/KXm The sorption data were described using the Freundlich equation as: 1 = 𝐾𝐾𝐾𝐾 𝑋𝑋 Where: 𝑛𝑛 K = Freundlich equilibrium constant C = Sorbent concentration in solution X = Amount adsorbed n = Freundlich exponent Data were fitted to the logarithmic form of the Freundlich equation = log + 1 A plot𝑙𝑙𝑙𝑙𝑙𝑙 of 𝑙𝑙logX vs.𝐾𝐾 logC𝑛𝑛 yields𝑙𝑙𝑙𝑙𝑙𝑙𝑙𝑙 a slope = 1/n and an intercept = logK -1 The soil sorption coefficient Kd(mL g ) can be expressed as:

𝑑𝑑 𝑋𝑋 The organic carbon (OC) normalized sorption𝐾𝐾 coefficient (Koc) can be determined as: 𝐶𝐶 100 =

𝐾𝐾𝑜𝑜𝑜𝑜 𝐾𝐾𝑑𝑑 𝑂𝑂𝑂𝑂 4.3 Results and discussion 4.3.1 Sorption kinetics

Figure 4.1 shows the sorption kinetics of MAP and MDMA in three test soils. The sorption processes of both target compounds were mostly achieved within the first 6 h of reaction, and then became more gradual until the equilibrium point (12 h). The results showed that a rapid initial sorption occurred followed by a slow approach to equilibrium. In general, organic contaminants tend to adsorb more quickly into natural sorbents in the initial stage, and then gradually reach equilibrium (Delle Site, 2001). The sorption rate of the CW soil for MAP was a little faster than that of the ML and W soils, and their sorption capacity at 12 h was 6.8, 5.0 and 5.1 µg mL-1, respectively. Almost the same sorption trend was observed for MDMA. This may be mainly attributed to the higher amount of OC in the CW soil. The present findings corroborate those of Rojas et al. (2013) and Studzińska et al. (2008) regarding the influence of organic matter content on sorption rate and various organic pollutants’ sorption capacity.

Figure 4.1 Sorption kinetics of MAP & MDMA in 3 experimental soils

4.3.2 Sorption isotherms

The sorption isotherms for MAP and MDMA alone and in mixtures are shown in Figure 4.2. The sorption isotherms of these target compounds were tested for five initial concentrations ranging from 5 to 60 μg-mL−1. Results revealed that the sorption of the target compounds increased steadily as concentrations increased (Figure 4.2). As per the r2 (˃ 0.74) values, the results fits well with Freundlich isotherms than Longmuir isotherms (Table 4. 2). Similar results were reported by Pal et al. (2012) for three soils on the MAP by-product 1-(1′, 4′-cyclohexadienyl)-2- methylaminopropane (CMP). The present results are well described by Freundlich isotherms, indicating non-linear sorption of the target compounds. Rojas et al. (2013) and De Wilde et al. (2009) explained that this absence of linearity may be due to specific interaction between the

natural sorbent’s organic matter and polar groups of pollutants. The 1/n value in the Freundlich equation was used as a directory to measure the non-linearity (He et al., 2006). The 1/n values in this study fit the non-linear Freundlich equation (1/n, 0.54 - 0.85), when the 1/n values ˂ 1 indicate the Freundlich and ˃1 for the Langmuir isotherm sorption patterns (Voudrias et al., 2002).

Figure 4.2 A plot for the sorbed amount (μgg-1) of MAP & MDMA in 3 experimental soils as a function of equilibrium concentration (μgmL-1)

4.3.3 Sorption parameters

The K values in this study when compared to all the treatments revealed that higher values for MAP alone and in a mixture, followed by MDMA alone and in mixture (Table 4.2). The CW soil recorded maximum K values in all the treatments and the values decreased in the order of CW>ML>W soils. This result indicates that the CW soil with highest OC has higher sorption potential to MAP corroborating well with the report of Pal et al. (2012).

The sorption parameters (Kd, Koc and Gibb's energy) were calculated for MAP and MDMA alone and as mixture in three soils, and the results are presented in Table 4.3. -1 The sorption coefficient (Kd) values for the three soils ranged from 5.13 to 9.96 mLg and 5.39 to 9.73 mLg-1 for MAP alone and as a mixture while the corresponding values

for MDMA were 5.48 to 8.24 mLg-1 and 5.66 to 8.61 mLg-1, respectively. These results show that a moderate level of MAP and MDMA sorption to the test soils occurred. The sorption affinity of these two compounds alone and as a mixture to the experimental soils was as follows: MAP ˃ MAP mixture ˃ MDMA mixture ˃ MDMA. The organic

carbon normalised sorption coefficient (Koc) values that followed were similar to Kd -1 values except for a few exceptions. The higher and lower (437.06 & 175.64 mL-g ) Koc values were recorded for the MDMA mixture and single compounds in W and ML soils,

respectively. Overall the Koc values for the test soils ranged between: firstly, 203.85 and 264.89 mL-g-1 for MAP; and secondly, between 219.55 and 273.60 mL-g-1for MAP

mixtures. In the case of MDMA alone and as a mixture the Koc values ranged between 175.64 and 316.24 mL-g-1and between 181.41 and 437.06 mL-g-1, respectively. The Gibb’s energy (ΔG0- Kcal mol-1) was calculated using all these experimental data in the integrated forms of the Gibb's sorption equations. The (ΔG0) values irrespective of the treatments and test soils ranged between -3.06 to -3.60 Kcal mol-1. These close ranges of negative values of ΔG0 for all treatments and soils indicate that these sorption forces are strong enough to break the barriers during the sorption process. However, it is difficult to conclude with specificity, using only three experimental soils, the main force and interaction involved in these target compounds’ sorption.

Table 4.3 Summary of the sorption parameters of MAP & MDMA in 3 experimental soils (Kd, Koc & Gibb’s energy)

Chemicals ML soil CW soil W soil

Kd Koc ΔG° Kd Koc ΔG° Kd Koc ΔG° (mLg-1) (mLg-1) (Kcal (mLg-1) (mLg-1) (Kcal (mLg-1) (mLg-1) (Kcal mol-1) mol-1) mol-1) MAP 6.4 203.9 -3.2 9.9 264.9 -3.3 5.1 260.4 -3.3 MDMA 5.5 175.6 -3.1 8.2 219.2 -3.2 6.2 316.2 -3.4 MAP 6.9 219.6 -3.2 9.7 258.8 -3.3 5.4 273.6 -3.3 mixture MDMA 5.7 181.4 -3.1 8.6 228.9 -3.2 8.6 437.1 -3.6 mixture

Table 4.2 Summary of the sorption parameters of MAP & MDMA in 3 experimental soils

Chemicals Langmuir Freundlich ML soil CW soil W soil ML soil CW soil W soil K r2 K r2 K r2 K n r2 K n r2 K n r2 MAP 4.55 0.13 8.33 0.24 3.33 0.29 3.55 0.79 0.91 7.08 0.85 0.97 2.19 0.73 0.94 MDMA 1.25 0.21 3.45 0.30 2.86 0.23 1.51 0.54 0.74 1.82 0.58 0.88 1.70 0.65 0.87 MAP 5.26 0.17 7.58 0.46 1.60 0.22 3.98 0.80 0.94 6.31 0.81 0.99 1.23 0.56 0.79 mixture MDMA 2.49 0.13 4.00 0.27 3.23 0.16 2.09 0.72 0.76 2.45 0.63 0.88 2.57 0.71 0.81 mixture

4.3.4 Correlation of sorption coefficient (Kd) with major soil properties

In general, the quantity of freely dissolved portion of chemicals in water is readily available for soil sorption (Wen et al. 2012). However, the sorption potential and rate of sorption of a soil is determined by some key soil properties such as soil OC, clay, dissolved organic matter (DOC), CEC and pH (Cao et al., 2008; Chefetz et al., 2008). In this study the following soil properties (OC˃ DOC˃ clay) were positively correlated with the sorption

coefficient (Kd) of MAP while clay˃ OC˃ DOC were positively correlated in a descending order for MDMA (Table 4.4). At the same time the following soil characters (CEC˃ EC˃ sand) were negatively correlated with both MAP and MDMA. A very similar trend was observed in the target compounds’ mixture of treatments. This study’s results suggest that the sorption of MAP and MDMA in these three experimental soils was governed by soil OC, clay and DOC. The soil’s organic matter and clay particles significantly promote the sorption of organic chemicals (Zhang et al., 2014; Charles et al., 2006). The sorption of organic pollutants by soil increased as soil OC also increased, which mostly influenced positively the soil sorption capacity (Rojas et al., 2013; Sun et al., 2008b). The soil DOC also influences sorption of organic pollutants, and the level of sorption depends on the source and amount of DOC in soil (Gao et al., 2007; Sun et al., 2008a). Despite the OC, clay and DOC showing a positive influence on MAP and MDMA sorption in the three test soils, the extent of specific interaction or mechanisms involved needs to be examined further using more soil types.

Table 4.4 Correlation of sorption coefficient (Kd) of MAP & MDMA with soil properties.

pH Electrical Cation Organic Dissolved Sand Silt Clay BET

conductivity exchange carbon Organic

capacity carbon

MAP -0.06 -0.59 -0.76 0.92* 0.91 -0.36 -0.01 0.90 -0.09

MDMA 0.44 -0.11 -0.34 0.61 0.58 -0.77 0.49 0.99* 0.41

*Significant at p ≤ 0.05.

4.3.5 Desorption

The percent desorption of the target compounds in three soils is presented in Figure 4.3. In all treatment cases, the percent desorption was initially rapid, but gradually declined over time. The target compounds were desorbed in soils in the following order: W> ML˃ CW. Of these treatments desorption occurred in the following descending order: MAP ˃ MAP mixture ˃ MDMA ˃ MDMA mixture. Overall, MAP alone is very likely to leach in all test soils, and among the test soils the CW soil showed the least percentage of leaching -1 throughout the treatments. The CW soil also had more Kd (8.24 – 9.96 mLg ), OC (3.76 %) and moderate clay (19.82%), which may have influenced these compounds becoming less mobile in soil. A similar finding was reported by Pal et al. (2012) who concluded that

Kd, OC and clay were linked to lower desorption of CMP in Sturt soil.

Figure 4.3 Desorption pattern of MAP & MDMA in 3 experimental soils

4.4 Conclusion In this study, three soils’ sorption and desorption patterns of MAP and MDMA alone and as a mixture were tested. The results indicated that: 1) MAP alone showed higher sorption potential as well as desorption in all experimental soils and treatments (Table 4.5);

2) the sorption coefficient (Kd) of MAP was positively correlated with OC, DOC and clay while for MDMA it was clay, OC and DOC; 3) CEC, EC and sand were negatively

correlated with sorption coefficient (Kd) of all treatment cases; 3) the CW soil showed higher percent of sorption whereas the W soil was highly prone to leaching of the target compounds (Table 4.5). Finally, after five desorption cycles the CW soil retained the largest amount of MDMA mixture and MDMA (112.67 and 95.73 µg g-1). At the same time the W soil retained the least amount of MAP mixture and MAP (24.16 and 26.30 µg g-1) (Table 4.5), respectively. These results could give an insight into the sorption and desorption patterns and act as an information tool for the transport, fate and risk assessment of MAP and MDMA in the environment.

Table 4.5 Amount of chemical adsorbed and desorbed by soil from MAP (53.6 µgmL-1) and MDMA 54.7 µgmL-1) test solution.

Chemicals Soils Chemical adsorbed Chemical desorbed

Chemical % % desorption of Chemical adsorbed by soil adsorption total chemical remaining in (µg g-1) of total adsorbed soil after five chemical desorption added (µg g-1) MAP ML 143.8 17.9 66.7 47.9 CW 181.5 22.6 58.1 76.1 W 139.9 17.4 81.2 26.3 MAP ML 150.7 18.7 50.3 74.6 mixture CW 183.4 22.8 53.6 85.1 W 145.4 18.1 83.4 24.2 MDMA ML 142.3 17.3 60.2 56.7 CW 179.3 21.9 46.6 95.7 W 153.5 18.7 75.2 38.1 MDMA ML 134.7 16.4 60.2 53.6 mixture CW 179.3 21.9 37.2 112.7 W 151.0 18.4 79.3 31.3

4.5 References ACC (Australian Crime Commission), Illicit drug data report 2011-12. http://www.crimecommission.gov.au/publications/illicit-drug-data-report/illicit-drug- data-report-2011-12 [Accessed on 10th February 2014]. Cao J, Guo H, Zhu HM, Jiang L, Yang H. Effects of SOM, surfactant and pH on the sorption– desorption and mobility of prometryne in soils. Chemosphere 2008; 70(11):2127-2134. Charles SM, Li H, Teppen BJ, Boyd SA. Quantifying the availability of clay surfaces in soils for adsorption of nitrocyanobenzene and diuron. Environ Sci Techno 2006; 40(24): 7751-7756. Chefetz B, Mualem T, Ben-Ari J. Sorption and mobility of pharmaceutical compounds in soil irrigated with reclaimed wastewater. Chemosphere 2008; 73(8):1335-1343. De Wilde T, Spanoghe P, Ryckeboer J, Jaeken P, Springael D. Sorption characteristics of pesticides on matrix substrates used in biopurification systems. Chemosphere 2009; 75(1):100-108. Delle Site A. Factors affecting sorption of organic compounds in natural sorbent/water systems and sorption coefficients for selected pollutants. A review, J Phys Chemi Refere Data 2001; 30(1):187-439. Gao Y, Xiong W, Ling W, Wang X, Li Q. Impact of exotic and inherent dissolved organic matter on sorption of phenanthrene by soils. J Hazard Mater 2007; 140(1):138-144. He Y, Xu J, Wang H, Ma Z, Chen J. Detailed sorption isotherms of pentachlorophenol on soils and its correlation with soil properties. Environ Res 2006; 101(3):362-372. Pal R, Megharaj M, Kirkbride KP, Naidu R. Illicit drugs and the environment- A review. Sci Total Environ 2013; 463-464:1079-1092. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor, and by-products in soil. Chemosphere 2011; 85:1002-1009. Pal R, Megharaj M, Kirkbride KP, Naidu R. Fate of 1-(1′, 4′-cyclohexadienyl)-2- methylaminopropane (CMP) in soil: Route-specific by-product in the clandestine manufacture of methamphetamine. Sci Total Environ 2012; 416:394-399. Rojas R, Morillo J, Usero J, Delgado-Moreno L, Gan J. Enhancing soil sorption capacity of an agricultural soil by addition of three different organic wastes. Sci Total Environ 2013; 458-460:614-623. Studzińska S, Sprynskyy M, Buszewski B. Study of sorption kinetics of some ionic liquids on different soil types. Chemosphere 2008; 71(11):2121-2128. Sun Z, Mao L, Xian Q, Yu Y, Li H, Yu H. Effects of dissolved organic matter from sewage sludge on sorption of tetrabromobisphenol A by soils. J Environ Sci 2008a; 20(9): 1075-1081.

Sun Z, Yu Y, Mao L, Feng Z, Yu H. Sorption behavior of tetrabromobisphenol A in two soils with different characteristics. J Hazard Mater 2008b; 160(2-3):456-461. UNODC (United Nations Office on Drugs and Crime). World Drug Report 2008. United Nations publication. http://www.unodc.org/documents/wdr/WDR_2008/WDR_2008. [Accessed on October 25, 2013]. Voudrias E, Fytianos K, Bozani E. Sorption–desorption isotherms of dyes from aqueous solutions and wastewaters with different sorbent materials. Global Nest 2002; 4(1):75- 83. Wen Y, Su LM, Qin WC, Fu L, He J, Zhao YH. Linear and non-linear relationships between soil sorption and hydrophobicity model, validation and influencing factors. Chemosphere 2012; 86(6):634-640. Zhang D, Hou L, Zhu D, Chen W. Synergistic role of different soil components in slow sorption kinetics of polar organic contaminants. Environ Pollut 2014; 184:123-130.

CHAPTER 5

Degradation of cocaine in soils and its adverse effects on earthworms (Eisenia fetida)

Abstract

The illegal production and consumption of cocaine results in release of substantial amount of residue into the environment. Current study presents the first systematic investigation of cocaine degradation pattern and its adverse effects on earthworms in soil. The persistence of cocaine was studied in vitro for the duration of 120 days in three different soils under both non-sterile and sterile conditions. The degradation results showed that cocaine degrades very rapidly in a non-sterile condition in all three soils (half-life between 2.2 and 3.9 days) compared to sterile condition (half-life between 40.8 and 54.1 days). Degradation of cocaine in the soil resulted in the formation of products such as benzoylecogonine (BE) and ecgonine methyl ester (EME). BE was relatively stable for a period of time in non-sterile soil compared to cocaine. The chronic toxicity of cocaine to earthworms were evaluated after 28 days of soil exposure. Significant weight loss and lipids reduction were prominent at ˃ 25 mg kg-1 of exposure, suggesting that interference in growth and energy reserve in cocaine toxicity. Significant increase in DNA damage, olive tail movement, and lipid peroxidation were found at concentrations ˃ 1 mg kg-1 and significant impact on total antioxidant capacity was evident (˃ 25 mg kg-1). This study confirmed that adverse effects on DNA and antioxidants enzymes were due to cocaine-induced oxidative stress.

Keywords: Cocaine; Illicit drugs; Soil; Toxicity; Earthworms

5.1 Introduction Illicit drugs represent an important group of emerging organic micropollutants. In environmental systems (wastewater, surface waters, sewage sludge and atmospheric air), several types of illicit drugs have been detected and reported worldwide. These compounds are discharged into the environment mainly through wastewater treatment plants and illegal disposal of wastes from clandestine laboratories (Pal et al., 2013). The commonly detected compounds in the environmental systems include cocaine, cannabis, amphetamines, opiates, and lysergic acid diethylamide (Janusz et al., 2003; UNODC, 2008). Globally, the illicit use of cocaine is a major concern and only second to cannabis and amphetamine-type stimulants. 82

Consequently, the environmental impact of these potentially toxic substances is being recognized as a critical issue. In Australia over the last decade, a relatively stable population (11%) has been reported using cocaine while in recent years (2012-13) a highest number of cocaine seizures and arrests were reported (ACC, 2012-13). The fate and toxicity of cocaine in the environment, particularly in soils, still remains poorly unstated. In general, the chemicals used for manufacturing of illicit drugs in clandestine laboratories including the precursors and by-products as well as the parent compounds are often buried in soil or disposed of in sinks or toilets (Janusz et al., 2003; Scott et al., 2003). Our research group has investigated the degradation pattern of the following illicit drugs, precursors and associated by-products namely: methylamphetamine sulphate (MAS), methamphetamine (MAP), 3, 4.-methylendioxymethamphetamine (MDMA), phenyl-2- propanone (P2P), pseudoephedrine (PSE), N-formylmethylamphetamine and 1-benzyl-3- methylnaphthalene in different soils. The results revealed that P2P and PSE were rapidly degraded in all the experimental soils while the MDMA remained moderately stable, whereas MAS and MAP were highly persistent in soil over a period of 6 weeks and 1 year, respectively. Yet, the degradation pattern of cocaine in soil is still unknown. In recent years, environmental scientists have recognized the growing problem that cocaine and its metabolites represent, and it is necessary to investigate their fate and toxicity in the environment. Most previous studies on illicit drugs focussed on analytical techniques, chemical impurity profiling, sewage epidemiology and mammal toxicity (Pal et al., 2011; Pal et al., 2013). To date no information is available in the literature on cocaine degradation in soil and its terrestrial toxicity. To assess the environmental toxicity of cocaine, earthworms (Eisenia fetida) were selected because they live in close contact with the soil through their skin or gut and process large quantities of soil. These ecologically relevant characteristics of earthworms enable us to investigate soil contaminants like cocaine’s impact on life parameters and cell damage. Furthermore, E. fetida is very common, easy to handle and culture in laboratory conditions. Since most illicit drug wastes reach the soil in a variety of ways, the earthworms are the most relevant organisms for testing the environmental impact of cocaine. The present research was designed to evaluate the degradation pattern of cocaine in soils using different physico- chemical properties (Table 4.1) and its adverse effects on earthworms.

5.2 Materials and methods 5.2.1 Chemicals Refer chapter 3.2.1

5.2.2 Test organisms

E. fetida was selected as a test species to evaluate the toxicity of cocaine in soil. The initial batch of earthworms was purchased from Bunning’s Warehouse, Parafield Gardens, South Australia to establish the laboratory culture. These worms were maintained at 20 ± 2 °C, 72 ± 8% humidity and a 16:8 light/dark cycle, and fed with oatmeal (Magic Worm Food, Magic Products). All the tests in this study utilized adult worms weighing between 350 to 600 mg with a well-developed clitellum.

5.2.3 Soils In this study, we used CW soil for cocaine toxicity and all three soils (ML, CW and W) were subjected to the degradation test. Refer chapter 4.2.2 for soil collection and preparation. The physico-chemical properties of the soils were analysed using standard analytical methods and the results are provided in Table 4.1.

5.2.4 Cocaine degradation in soil Cocaine degradation was studied under both non-sterile and sterile conditions in three physico-chemically varied soils, as suggested by Pal et al. (2011). The experimental soils (5 g) were weighed into individual amber coloured glass vials fitted with Teflon lined solid screw caps. The soils in these vials were autoclaved at 121 °C for 20 min on three consecutive days for sterile degradation and subsequently the sterile conditions were maintained throughout the study period. The test soils (sterile and non-sterile) were spiked with 100 mg kg-1 of cocaine in all individual vials separately. The stock solutions (2 g L-1) were prepared in water and spiked in non-sterile soil while in sterile soil, the filter sterile (0.45 µm filters) stock solutions were spiked aseptically in laminar airflow cabinet. All the spiked soils were uniformly mixed by vortexing (10 sec). Control soils were also maintained for both non-sterile and sterile conditions without drug. Non-sterile soil vials were aerated aseptically within a laminar airflow every week. The moisture contents of the soils (both in non-sterile and sterile) were maintained by the aseptic addition of sterile deionised water. All the experimental treatments were conducted in duplicate. The concentrations of cocaine were monitored at intervals for up to 120 days. Cocaine was extracted from soil using 40 mL of extraction solvents (chloroform: acetonitrile: methanol: acetic acid - 80:10:9:1) in two steps, 20 mL for each step. The supernatants were centrifuged and filtered using 0.22 µm Teflon filters. The combined 40 mL of supernatant was dried under a nitrogen stream and reconstituted with methanol for analysis.

5.2.5 Toxicity assay Cocaine toxicity on E. fetida was studied according to the OECD guidelines (1984). All the tests were conducted in triplicate under controlled environment. The target compound concentration in the soil ranged between 1 and 100 mg kg-1. Each replicate consisted of 500 g of soil that was maintained with 70% of soil moisture. The soil without chemicals served as the control. Ten adult worms (E. fetida) were released into each test container and the soil moisture was maintained. Before and at the end of the experiments the earthworms were washed and placed on wet filter paper in a ventilated container to empty their guts for 24 - 48 h. Depurated earthworms’ weight was recorded before exposure to soil and at the end of the test. The earthworms’ behaviour and mortality were monitored and recorded daily during the test period.

5.2.6 Lipids estimation Cocaine exposed (28 days) earthworm lipid content was quantified by Folch method as described by Pérez-Palacios et al. (2008). Depurated earthworms (5 g) were mixed with 100 ml of chloroform: methanol (2:1, v/v). The mixture was homogenized in a tissue homogenizer, centrifuged at 3000 g, 10 min and then filtered. The supernatant was mixed with 5 ml of distilled water and vortexed. Subsequently the mixture was centrifuged at 3000 g for 10 min and the upper aqueous phase was removed. The lower phase (chloroformic) was filtered through anhydrous sodium sulphate and evaporated to determine the lipid amount gravimetrically.

5.2.7 Total antioxidant capacity Total antioxidant capacity was carried out with the kit obtained from BioVision. Basically, antioxidants play a vital role in preventing the formation and scavenging of free radicals and other toxic oxidizing species. There are three species of antioxidants: enzyme systems (GSH reductase, catalase, peroxidase, etc.); small molecules (ascorbate, uric acid, GSH, vitamin E, etc.); and proteins (albumin, transferrin, etc.). All enzymes vary in their reducing capacity; here in the kit Trolox is used to standardize antioxidants and other antioxidants being measured in Trolox equivalents. Measurement of the combined no enzymatic antioxidant capacity of biological fluids indicates to us its overall ability to counteract reactive oxygen species. Assay procedure includes the addition of Cu2+ working solution to all standard and sample wells. Later they were closed and left undisturbed for 1.5 h under room temperature. After incubation microplates were read at 570 nm absorbance using a plate reader.

Calculation

Sample antioxidant capacity = Sa/Sv = n mol/µl or mM Trolox equivalent

Sa is the sample amount read from the standard curve.

Sv is the undiluted sample volume added to the wells.

5.2.8 Lipid peroxidation Lipid peroxidation in earthworm tissue was carried out with the help of a lipid peroxidation kit from BioVision (Catalogue #K739-100: 100 assays). In the initial sample preparation phase, clean earthworms’ tissue samples weighted around 10 mg had been taken and homogenized with 300 µl of MDA lysis buffer provided in the kit (with 3 µl BHT(100X). Then they were centrifuged for 13,000 g for about 10 min in order to remove the insoluble fractions. Later 200 µl of supernatant was placed in the microcentrifuge tube. Standard curve was derived using MDA standards provided with the kit. 600 µl of TBA was added to both the samples and standards, which were then incubated at 95 °C for 60 min. Following incubation they were read at 532 nm in a microplate reader.

Calculation = × 4 × n mol/mg 𝐴𝐴 Where: 𝐶𝐶 �𝑚𝑚𝑚𝑚� 𝐷𝐷 A: Sample MDA amount from the standard curve (in n mol) mg: Original tissue amount (10mg) 4: Correction for using 200µl of the 800µl reaction mix D: Dilution factor

5.2.9 Comet assay Earthworms’ DNA damage and tail movements were evaluated by the use of alkaline single cell gel electrophoresis or comet assay, as suggested by Singh et al. (1988) with minor modifications. E. fetida coelomocytes were obtained by placing the animals in 1 ml of extrusion fluids for 3 min which induce secretion of coelomocytes containing coelomic fluids (Eyambe et al., 1991). The cells containing extrusion ﬂuid were centrifuged and the supernatants were removed followed by three micro centrifugations using 100 mM PBS for 3 min at 8000 g. Fifty micro litres of the cells were mixed with 75 µL of low melting agarose and mixed thoroughly by pipetting. Comet assay slides were coated with 50 µL of cells-agarose suspension and they were allowed to solidify for 5 min at a temperature of 4 °C. The alkaline comet assay was done according to the manufacturer’s instructions (Trevigen comet assay protocol, 8405 Helgerman Ct.). At the end of the assay, the slides were examined using a fluorescence microscope

(Olympus BX41) at 10X magnification. DNA damage was expressed as the tail and olive tail movement using an image analysis computerized method developed by CometScore software (TriTek Corp., USA).

5.2.10 Chemical analysis Cocaine and its metabolites (benzoylecogonine (BE) and ecgonine methyl ester (EME)) were determined by HPLC-MS (Refer chapter 3.2.3). The limits of quantitation (LOQ) was set 6 ng L-1 for cocaine, after established the concentration of the analyte in distilled water that gives rise to peak height with an S/N of 10. In this extraction method the analyte average recoveries was 75 ± 1.3 to 97 ± 1.2 when n = 3. Method repeatability, evaluated as the relative standard deviation (RSD) of the replicate analysis of spiked (50 ng L-1) water samples was also satisfactory, with RSD values between 1.5 to 5.4%.

5.3 Results and discussion 5.3.1 Cocaine degradation (non-sterile and sterile) The changes in the concentrations of cocaine throughout the incubation period for all three soils under both non-sterile and sterile conditions are shown in Figures 5.1 and 5.2. The initial degradation of cocaine was almost similar (up to 3 days) in both non-sterile and sterile conditions in all three soils. In non-sterile conditions, cocaine was rapidly degraded and over 7 days the loss (in percentage terms) of cocaine was recorded at 93.3 ± 0.4, 46.1 ± 0.7, and 60.0 ± 0.7 for ML, CW and W soils, respectively. Cocaine’s original concentration had virtually fallen to zero in ML soil after 15 days in non-sterile conditions, while the concentrations in CW and W soils declined to 6.8 ± 0.4% and 7.6 ± 0.2%, respectively. These results are comparable to Gheorge et al. (2008) and Postigo et al. (2008) who found that cocaine’s degradation (˃75%) in aqueous matrices at 20 °C in 5 days was dramatic. In non-sterile conditions such rapid decline of cocaine in soils may be predominantly caused by microbial activity in addition to other processes (chemical hydrolysis and photolysis), which agrees with the faster degradation of cocaine observed in natural aqueous matrices compare to sterile medium (Postigo et al., 2008). A complete degradation of cocaine (within 15 days) in ML soil compared to CW and W soils might be related to early adaption or abundance of efficient microbes, which could have utilised cocaine as source of carbon. Cocaine degradation products (BE and EME) began to appear in all three soils’ extracts at day 3, and progressively increased the peak area to a maximum level at days 7 and 15 (visual observation). Degradation product EME began to disappear over the next few days while BE remained stable for a period of time. BE peaks were observed in soil extraction between 3 and 45 days, but the peak areas began to fall after 45 days; during the experiment 87

period no compounds were found to accumulate. This result suggests that the microbial processes contributes to be a major factor involved when compared to the abiotic process, because in non-sterile conditions microorganisms completely transformed cocaine into metabolites, which in turn are also consumed. Degradation products (BE and EME) were identified on the basis of their mass spectral fragmentation pattern, then positively identified by HPLC-MS analysis of soil extract spiked with authentic compounds.

Figure 5.1 Non-sterile degradation of cocaine in three test soils

Figure 5.2 Sterile degradation of cocaine in three test soils

Under sterile conditions, loss of cocaine was also observed (in percentage terms) at 23.3 ± 1.2, 33.3 ± 0.8, and 33.3 ± 0.3, for ML, CW and W soils over 15 days, respectively. However, the degradation of cocaine between 15 and 45 days was very slow, and subsequently cocaine remained practically constant (90 days) in the ML and CW soils while a slight variation was observed in the W soil. In these periods degradation products (BE and EME) peaks were observed with constant peak areas. It is clearly suggested that the 88

degradation of cocaine in sterile soil could be due to two processes, either: firstly, biotic degradation by microorganisms that survived sterilisation; or secondly, abiotic process such as chemical hydrolysis or photolysis. All the materials, chemicals and soils used in the sterile degradation study were thoroughly sterilized and every step was very strictly executed to avoid microbial activity. The incubation was done under dark throughout the experimental period which suggests that chemical hydrolysis may be the major factor for sterile soil degradation. Thus, degradation of cocaine in soil may be influenced by the level of microbial activity, chemical hydrolysis and soil’s physico-chemical properties.

Table 5.1 Regression equation, rate constant (k), and half-life (t1/2) values for the degradation

of cocaine under non-sterile and sterile conditions

Non-sterile Sterile

soil

Regression equation t1/2 d Regression equation t1/2 d

Mount Lofty

(ML) y = -0.3225x + 4.6851 2.2 y = -0.0129x + 4.4769 53.7

Crafers West

(CW) y = -0.183x + 4.89 3.8 y = -0.0128x + 4.4725 54.1

Wallaroo (W) y = -0.1762x + 4.8144 3.9 y = -0.0168x + 4.417 40.8

The regression equations, regression coefficient (r2), rate constant (k-1), and half-life

(t½) values to describe the degradation of cocaine in both non-sterile and sterile conditions are presented in Table 5.1. Our data were fitted to simple regression equations considering first order reaction, where y is concentration and x is time. The half-life values were calculated from the best fit lines of the logarithm of residual concentrations vs. time elapsed in the incubation period. The half-life (t½) value for the non-sterile soil was predicted using the regression equation, and the values were 2.2, 3.8 and 3.9, respectively for ML, CW and W soil, while the corresponding values in sterile soil were 53.7, 54.1 and 40.8. Thus the results showed that cocaine had comparatively high half-value in CW soil under both non-sterile and sterile conditions which could be due to the neutral pH (7.2) and high organic carbon (3.8%) and dissolved organic carbon (14.7 µg ml-1) (Table 4.1).

5.3.2 Impact of cocaine on earthworms’ weight and total lipids content The weight loss and lipid reduction in earthworms due to cocaine exposure in a soil over 28 days are presented in Figure 5.3. Smaller cocaine concentrations, i.e. 1 and 5 mg kg- 1, did not affect the earthworms’ weight while concentrations ˃ 25 mg kg-1 did significantly reduce their weight compared to the control. In the control, a weight gain of 0.7% was recorded whereas at a higher concentration (100 mg kg-1) more than half of the original weight (51%) decreased. LC50 value for cocaine could not be calculated since no mortality was recorded throughout the concentration levels. Weight changes in earthworms due to chemical exposure constitute a sensitive indicator of chemicals’ toxicity in soil (Yasmin and D'Souza, 2010). Helling et al. (2000) observed a weight gain in control worms which agrees with our present study’s results. The main reason for the weight loss in cocaine exposure to earthworms is not clear. However, a similar effect was recorded in human beings and animals when cocaine was administered repeatedly (Shimosato et al., 1994). Furthermore, cocaine was reported in mammals a significant behavioural sensitization (Karler et al., 1993), sleep deprivation (Berro et al., 2014); and loss of appetite (Cooper and Van der Hoek, 1993; Ersche et al., 2013), the above effects may also be associated with cocaine-induced weight loss in earthworms. A similar type of weight loss outcome was reported for earthworms exposed to pesticides and metals (Espinoza‐Navarro and Bustos‐Obregón, 2005; Klok et al., 2006). In general, an organism’s responses to toxic chemicals may be evidenced by mortality or by significant physiological stresses which lead to several side effects including weight loss (Frampton et al., 2006). Cocaine has also been documented as initiating protein alteration, cyto-genotoxicity, neurotoxicity, oxidative stress and impairments in cognitive functions in various organisms (Morrow et al., 2002; Portugal-Cohen et al., 2010; Binelli et al., 2012 and 2013; Parolini et al., 2013). This may also be attributed in this present study to earthworms’ weight loss.

Figure 5.3 Effect of cocaine on earthworms’ weight and lipids after 28 days soil exposure. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. Different letters show statistically difference at p < 0.05 (ANOVA)

Furthermore, the present result revealed a significant reduction of lipids in cocaine exposed earthworms. The effects of cocaine on earthworms’ lipids were concentration dependent. Smaller concentrations of cocaine, 1 and 5 mg kg-1, did not affect the earthworms’ lipids content while concentrations ˃ 25 mg kg -1 significantly reduced the lipids content compared to the control (Figure 5.3). Although the mechanisms for earthworm’ lipids reduction are not known, it is clear from the results that lipids reduction can explain the weight loss in earthworms. In this study, earthworms exhibited certain behavioural and morphological changes such as coiling, poor burrowing, slower response to external stimuli and skin dehydration which may also be a possible reason for the above reduction in earthworms. Several studies have reported that humans and animals exposed to cocaine induced various behavioural and pathological conditions (Lange and Hillis, 2001; Fillmore and Rush, 2002). A recent study revealed that weight loss in cocaine-dependent individuals was associated with a significant reduction in fat mass level (Ersche et al., 2013) which reflects the fundamental perturbations of cocaine in fat regulation. Decreases of energy reserves (lipids, protein and carbohydrates) can be used as a biomarker for soil microorganisms including earthworms to assess the level of soil contamination (Beaumelle et al., 2014). The above findings support our finding that cocaine-induced weight loss and lipids reduction in earthworms can serve as a biomarker for cocaine soil contaminations.

5.3.3 Effect of cocaine on antioxidant capacity of earthworms This study examined the effect of cocaine on earthworms’ antioxidant responses. The total antioxidant capacity in E. fetida exposed to cocaine is presented in Figure 5.4. The result shows that total antioxidant content gradually increased in smaller concentrations and peaked at 25 mg kg-1 of cocaine exposed in neutral soil. The antioxidant capacity began to decline at larger concentrations, i.e. 50 mg kg-1 and 100 mg kg-1. The results presented here 91

demonstrated the cocaine exposure to earthworms significantly affects the total antioxidant capacity. “Oxidative stress” is a disturbance in the balance between the production of reactive oxygen species (ROS) and antioxidant defences (Li et al., 1999). Any changes in an organism’s defence system lead to various physiological dysfunctions in most of the cellular compounds such as DNA, proteins, and lipids (Nel et al., 2006). Antioxidant enzymes in an organism play a vital part in balancing the defence system to combat the oxidative stress. Antioxidants play an important role in preventing the formation and scavenging of free radicals and other potentially toxic oxidizing species (Peake et al., 2007). These antioxidants are represented in two groups, specifically non-enzymatic (reduced glutathione and alpha- tocopherol) and detoxification enzymes, for example superoxide dismutase and glutathione peroxidase (Yoganathan et al., 1989). Of these, reduced glutathione is critical in the defence system of an organism; it has an unusual linkage between the amine group of cysteines and carboxyl group of the glutamate side chain. The detoxification process using reduced glutathione mainly involves the bond between the sulphydryl group with carbon or chlorine atoms of the pollutants and their metabolites (Zelikoff et al., 1996; Xu et al., 2013). Several studies have shown that cocaine exposure leads to depleted reduced glutathione and a decrease in total scavenging capacity in animals (Poon et al., 2007; Portugal-Cohen et al., 2010).

Figure 5.4 Effect of cocaine on total antioxidant capacity of earthworms after 28 days exposure in soil.

In our results variation in the total antioxidant capacity in earthworms (Figure 5.4) may be related to the depletion of glutathione at higher concentrations. Thus, reduction of

glutathione in earthworms encourages cocaine to escape from the defence system, in turn leading to the formation of free radicals and an oxidative stress condition. Binelli et al. (2012) observed cyto-genotoxicity effects in Dreissena polymorpha exposed to environmental concentrations of cocaine which was due to oxidative stress. Cocaine exposure to rat hepatocytes induced apoptosis as a result of increased oxidative stress (Díez-Fernández et al., 1999). Parolini et al. (2013) in their experiments on D. polymorpha exposed to cocaine metabolites BE (0.5 ng L-1 and 1 ng L-1), concluded that oxidative stress biomarkers such as catalase (CAT), superoxide dismutase (SOD), glutathione peroxide (GPx) and glutathione S- transferase activity increased in low concentrations (0.5 ng L-1) and decreased in high concentrations (1 ng L-1). Their finding agrees with ours concerning the earthworms. In addition EME induced a variation in D. polymorpha’s antioxidant enzymes (SOD, CAT, GPx and GST) at 0.15 and 0.5 ng L-1 which is linked to the induction of oxidative stress (Parolini and Binelli, 2013). This outcome supports that concluded in this present study, specifically that cocaine induced oxidative damage in earthworms.

5.3.4 Effect of cocaine on lipid peroxidation of earthworms Lipid peroxidation is a widely accepted biomarker for measuring oxidative stress because it is responsible for inducing various kinds of cell damage in an organism (Lushchak et al., 2001; Dalton et al., 2002). Lipid peroxidation of cell membranes occurs as a result of the production of ROS which sparks the peroxidative chain reaction (Li et al., 1999), and is directly proportional to the amount of fatty acid metabolism that occurs in the body. MDA is a product of peroxidative decomposition of fatty acids which is used to measure lipid peroxidation in an organism. Consequently, measuring the amount of MDA provides an index of oxidative stress and lipid peroxidation (Halliwell and Chirico, 1993). For the lipid peroxidation assay, MDA was measured in earthworms’ tissue during the period of cocaine exposure. Result shows that more MDA was produced in treatments compared to the control (Figure 5.5). Generally, the production of MDA increased with the increase in the cocaine concentrations; in fact lipid peroxidation was highly significant between 5 and 100 mg kg-1. The results show that earthworms exposed to cocaine enhanced the process of lipid peroxidation, as measured by the increase in MDA content.

Figure 5.5 Effect of cocaine on earthworms’ lipid peroxidation after 28 days soil exposure.

Our results indicated the formation of MDA in all the treatments exposed to cocaine. This was consistent with the study conducted in rat testis, where MDA production increased due to chronic cocaine exposure (Li et al., 1999). Some recent findings revealed that environmentally relevant concentrations of cocaine and its metabolites (BE and EME) exposed to a freshwater biological model have shown significant cyto-genotoxicity; this is a possible implication of oxidative stress (Binelli et al., 2012; Binelli et al., 2013; Parolini and Binelli, 2013; Parolini et al., 2013). Lipid peroxidation is one of the main processes induced by oxidative stress in an organism (Ahmad et al., 2004). Several studies on various biological models have demonstrated that cocaine exposure can cause oxidative stress (Li et al., 1999; Boess et al., 2000; Poon et al., 2007; Portugal-Cohen et al., 2010). Thus, the observed lipid peroxidation in our study could be due to oxidative stress induced by cocaine exposure to earthworms.

5.3.5 Cocaine induced DNA damage Earthworms exposed to cocaine for 28 days had their DNA damage evaluated by the data of percentage tail DNA and olive tail movement (Figure 5.6). Cocaine caused significant (p < 0.05) DNA damage and olive tail movement in low, medium and high concentrations compared the controls. This finding confirms that DNA damage and olive tail movement increased when the concentration increased, and the most DNA damage (16%) and olive tail movement (4%) was observed at 100 mg kg -1. A test comprising a smaller concentration (1 mg kg-1) also induced a significant percentage tail DNA and olive tail movement (4% and 5.5%) compared the controls. Figures (5.7a and b) show no or only minimal DNA damage in control, and DNA migration and damage as a result of strand breakage in cocaine exposed 94

earthworms’ cells. Several studies have been conducted on identifying the reason for DNA damage and most analyses reported that ROS accumulation in the tissue leads to genetic damage (Ames, 1983; Cooke et al., 2003). The above phenomena agree well with Binelli et al. (2012) who reported that environmental concentration of cocaine and its metabolites (BE and EME) induces oxidative stress, and as a result significant primary DNA damage, protein alteration and apoptosis to D. polymorpha. In addition the mitotic stimulation process can damage DNA. Stefanidou et al. (2002) observed the occurrence of mitogenic effects on protozoan (Tetrahymena pyriformis) upon exposure to cocaine products. Generally, mitogenic stimulation is harmful to cells and increases the DNA content. Cocaine’s genotoxic influence has been well documented in several organisms such as mouse oocytes (Combelles et al., 2000), human foetuses (Meyer and Zhang, 2009), protozoan T. pyriformis (Stefanidou et al., 2002), and D. polymorpha haemocytes (Binelli et al., 2012). In our study, DNA damage to earthworms could be due to one of the above mechanisms, which requires a detailed investigation to make this conclusive. However, measuring the genotoxic effects of cocaine employing a genotoxicity biomarker such as comet assay in earthworms could prove useful for monitoring the terrestrial toxicity of illicit substances.

Figure 5.6 DNA damage induced by cocaine in E. fetida. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA & olive tail movement - ANOVA, p< 0.05)

Figure 5.7 DNA damage in E. fetida exposed to cocaine (a). Control with no or minimal DNA damage (b). DNA damage in cocaine exposed E. fetida in soil.

(a) (b)

5.4 Conclusion The results obtained in this study clearly indicate the environmental degradation of cocaine is a complex process which depends on various factors. Overall, the degradation pattern of cocaine mostly relied on biotic and some abiotic processes in the soils. BE was relatively stable for a period of time in non-sterile soil compared to cocaine. The toxicity results obtained in this study revealed that cocaine exposure caused significant adverse effects on earthworms at ˃ 1mg kg-1 in soil, highlighting its possible hazard to terrestrial organisms. The experiment concerning chronic cocaine exposure (28 days) indicated that oxidative stress in exposed earthworms led to weight loss, reduced energy reserves, DNA damage and an imbalance in antioxidant enzymes. This suggests that cocaine- contaminated soil is dangerous to soil biota. Finally, cocaine degrades very rapidly in soil into BE and EME, leading to the conclusion that in reference to the potential risk faced by soil microorganisms, more studies are needed on cocaine metabolites’ adverse effects on soil biota.

5.5 References

ACC (Australian Crime Commission), Illicit drug data report 2012-13. http://www.crimecommission.gov.au/publications/illicit-drug-data-report/illicit-drug-data- report-2012-13 [Accessed on 10th May 2014]. Ahmad I, Pacheco M, Santos MA. Enzymatic and nonenzymatic antioxidants as an adaptation to phagocyte-induced damage in Anguilla anguilla L. following in situ harbour water exposure. Ecotoxicol Environ Saf 2004; 57:290-302. Ames BN. Dietary carcinogens and anticarcinogens oxygen radicals and degenerative diseases. Science 1983; 221(4617):1256-1264. Beaumelle L, Lamy I, Cheviron N, Hedde M. Is there a relationship between earthworm energy reserves and metal availability after exposure to field-contaminated soils? Environ Poll 2014; 191:182-189. Berro LF, Santos R, Hollais AW, Wuo-Silva R, Fukushiro DF, Mári-Kawamoto E, Costa JM, Trombin TF, Patti CL, Grapiglia SB, Tufik S, Andersen ML, Frussa-Filho R. Acute total sleep deprivation potentiates cocaine-induced hyperlocomotion in mice. Neurosci Lett 2014; 579:130-133. Binelli A, Marisa I, Fedorova M, Hoffmann R, Riva C. First evidence of protein profile alteration due to the main cocaine metabolite (benzoylecgonine) in a freshwater biological model. Aqu Toxicol 2013; 140-141:268-278. Binelli A, Pedriali A, Riva C, Parolini M. Illicit drugs as new environmental pollutants: cyto- genotoxic effects of cocaine on the biological model Dreissena polymorpha. Chemosphere 2012; 86:906-11. Boess F, Ndikum-Moffor FM, BoelsterlI UA, Roberts SM. Effects of cocaine and its oxidative metabolites on mitochondrial respiration and generation of reactive oxygen species. Biochem Pharmacol 2000; 60:615-623. Combelles CMH, Carabatsos MJ, London SN, Mailhes JB, Albertini DF. Centrosome-specific perturbations during in vitro maturation of mouse oocytes exposed to cocaine. Experi Cell Res 2000; 260:116-126. Cooke MS, Evans MD, Dizdaroglu M, Lunec J. Oxidative DNA damage: mechanisms, mutation, and disease. The FASEB J 2003; 17(10):1195-1214. Cooper SJ, Van Der Hoek GA. Cocaine: a microstructural analysis of its effects on feeding and associated behaviour in the rat. Brain Res 1993; 608:45-51. Dalton TP, Puga A, Shertzer HG. Induction of cellular oxidative stress by aryl hydrocarbon receptor activation'. Chemico-biol Interact 2002; 141(1):77-95.

Díez-Fernández C, Zaragoza A, Alvarez AM, Cascales M. Cocaine cytotoxicity in hepatocyte cultures from phenobarbital-induced rats: involvement of reactive oxygen species and expression of antioxidant defence systems. Biochem Pharmacol 1999; 58:797-805. Ersche KD, Stochl J, Woodward JM, Fletcher PC. The skinny on cocaine: Insights into eating behaviour and body weight in cocaine-dependent men. Appetite 2013; 71:75-80. Espinoza‐Navarro O, Bustos‐Obregón E. Effect of malathion on the male reproductive organs of earthworms, Eisenia foetida. Asian J Androl 2005; 7:97-101. Eyambe GS, Goven AJ, Fitzpatrick L, Venables BJ, Cooper EL. A non-invasive technique for sequential collection of earthworm (Lumbricus terrestris) leukocytes during subchronic immunotoxicity studies. Lab Animals 1991; 25:61-67. Fillmore MT, Rush CR. Impaired inhibitory control of behavior in chronic cocaine users. Drug Alcohol Dep 2002; 66:265-273. Frampton GK, Jänsch S, Scott‐Fordsmand JJ, Römbke J, Van Den Brink PJ. Effects of pesticides on soil invertebrates in laboratory studies: a review and analysis using species sensitivity distributions. Environ Toxicol Chem 2006; 25:2480-2489. Gheorghe A, van Nuijs A, Pecceu B, Bervoets L, Jorens PG, Blust R. Analysis of cocaine and its principal metabolites in waste and surface water using solid-phase extraction and liquid chromatography-ion trap tandem mass spectrometry. Anal Bioanal Chem 2008; 391:1309-19. Halliwell B, Chirico S. Lipid peroxidation: its mechanisms, measurement, and significance. The American J Clinic Nut 1993; 57:715-724. Helling B, Reinecke S, Reinecke A. Effects of the Fungicide Copper Oxychloride on the Growth and Reproduction of Eisenia fetida (Oligochaeta). Ecotoxicol Environ Saf 2000; 46:108-116. Janusz A, Kirkbride KP, Scott TL, Naidu R, Perkins MV, Megharaj M. Microbial degradation of illicit drugs, their precursors, and manufacturing by-products: implications for clandestine drug laboratory investigation and environmental assessment. Forensic Sci Int 2003; 134: 62-71. Karler R, Finnegan KT, Calder LD. Blockade of behavioral sensitization to cocaine and amphetamine by inhibitors of protein synthesis. Brain Res 1993; 603:19-24. Klok C, Van Der Hout A, Bodt J. Population growth and development of the earthworm Lumbricus rubellus in a polluted field soil: Possible consequences for the godwit (Limosalimosa). Environ Toxicol Chem 2006; 25:213-219. Lange RA, Hillis LD. Cardiovascular complications of cocaine use. New Engl J Med 2001; 345: 351-358.

Li H, Jiang Y, Rajpurkar A, Tefilli MV, Dunbar JC, Dhabuwala CB. Lipid peroxidation and antioxidant activities in rat testis after chronic cocaine administration. Urology 1999; 54: 925-928. Lushchak VI, Lushchak LP, Mota AA, Hermes-Lima M. Oxidative stress and antioxidant defenses in goldfish Carassius auratus during anoxia and reoxygenation, American J Physiol Regu Integ Compa Physiol 2001; 280(1):100-107. Meyer KD, Zhang L. Short- and long-term adverse effects of cocaine abuse during pregnancy on the heart development. Ther Adv Cardi Dise 2009; 3: 7-16. Morrow BA, Elsworth JD, Roth RH. Prenatal cocaine exposure disrupts non-spatial, short-term memory in adolescent and adult male rats. Behav Brain Res 2002; 129(1):217-223. Nel A, Xia T, Mädler L, Li N. Toxic potential of materials at the nanolevel. Science 2006; 311(5761):622-627. Organisation for Economic Co-operation and Development. Earthworm acute toxicity test, OECD Guideline 207; 1984, Paris, France. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor and by-products in soil. Chemosphere 2011; 85:1002-9. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Illicit drugs and the environment—A review. Sci Total Environ 2013; 463-464:1079-1092. Parolini M, Pedriali A, Riva C, Binelli A. Sub-lethal effects caused by the cocaine metabolite benzoylecgonine to the freshwater mussel Dreissena polymorpha. Sci Total Environ 2013; 444:43-50. Parolini M, Binelli A. Adverse effects induced by ecgonine methyl ester to the zebra mussel: A comparison with the benzoylecgonine. Environ Poll 2013; 182:371-378. Peake JM, Suzuki K, Coombes JS. The influence of antioxidant supplementation on markers of inflammation and the relationship to oxidative stress after exercise. The J Nut Biochem 2007; 18:357-371. Pérez-Palacios T, Ruiz J, Martín D, Muriel E, Antequera T. Comparison of different methods for total lipid quantification in meat and meat products. Food Chem 2008; 110:1025- 1029. Poon HF, Abdullah L, Mullan MA, Mullan MJ, Crawford FC. Cocaine-induced oxidative stress precedes cell death in human neuronal progenitor cells. Neurochem Int 2007; 50:69-73. PortugaL-Cohen M, Numa R, Yaka R, Kohen R. Cocaine induces oxidative damage to skin via xanthine oxidase and nitric oxide synthase. J Dermatol Sci 2010; 58:105-112. Postigo C, de Alda MJL, Barceló D. Analysis of drugs of abuse and their human metabolites in water by LC-MS2: a non-intrusive toll for drug abuse estimation at the community level. Trends Anal Chem 2008; 27:1053-69.

Scott TL, Janusz A, Perkins MV, Megharaj M, Naidu R, Kirkbride KP. Effect of amphetamine precursors and by-products on soil enzymes of two urban soils. Bull Environ Contam Toxicol 2003; 70:824-31. Shimosato K, Saito T, Marley RJ. Genotype-specific blockade of cocaine-induced weight loss by the protein synthesis inhibitor, anisomycin. Life Sci 1994; 55:293-299. Singh NP, McCoy MT, Tice RR, Schneider EL. A simple technique for quantitation of low levels of DNA damage in individual cells. Expe Cell Res 1988; 175:184-191. Stefanidou M, Chatziioannou A, Livaditou A, Rellaki A, Alevisopoulos G, Spiliopoulou H, Koutselinis A. DNA toxicity of cocaine hydrochloride and cocaine freebase by means of DNA image analysis on Tetrahymena pyriformis. Biol Pharm Bull 2002; 25:332-334. UNODC (United Nations Office on Drugs and Crime). Amphetamines and Ecstasy: Global ATS Assessment. United Nations Publication; 2008. http://www.unodc.org/documents/scientific/ATS/Global-ATS-Assessment-2008- Web.pdf [Accessed on October 15, 2014]. Xu D, Li C, Wen Y, Liu W. Antioxidant defence system responses and DNA damage of earthworms exposed to Perfluorooctane sulfonate (PFOS). Environ Poll 2013; 174:121- 127. Yasmin S, D'souza D. Effects of pesticides on the growth and reproduction of earthworm: a review. App Environ Soil Sci 2010; 1-9. Yoganathan T, Eskild W, Hansson V. Investigation of detoxification capacity of rat testicular germ cells and Sertoli cells. Free Rad Biol Med 1989; 7:355-359. Zelikoff JT, Wang W, Islam N, Twerdok LE, Curry M, Beaman J, Flescher E. Assays of reactive oxygen intermediates and antioxidant enzymes: potential biomarkers for predicting the effects of environmental pollution'. Tech AquToxicol 1996; 1:287-306.

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CHAPTER 6

Methamphetamine (MAP) toxicity to earthworms (Eisenia fetida) following soil exposure

Abstract We studied methamphetamine (MAP) toxicity to earthworm (Eisenia fetida) following exposure to spiked soil at 5 to 200 mg kg-1. Though no mortality was recorded even at the highest concentration, results showed significant weight loss in all treatments. However, chronic exposure in adults resulted alterations in earthworms’ morphology and behaviour, such as poor burrowing, coiling, skin dehydration and cuticle fragmentation. The reproduction capacity also declined at concentration especially above 20 mg kg -1. Exposure at higher concentrations of 50 - 200 mg kg-1 significantly reduced both cocoon production and juvenile hatching at 73.5 - 95.7% and 88.9 - 100%, respectively, indicating toxicity is higher to juvenile earthworms than those at the cocoon stage. Bioaccumulation of MAP was observed as low as 0.2 - 10.3 mg kg-1 in different experimental concentrations tested. Interestingly significant damage to earthworms’ DNA was recorded even at 5 mg kg-1 exposure concentration. Findings here suggest that MAP soil contamination is a threat to soil biota and the environment.

Keywords: Illicit drugs; Methamphetamine; Amphetamine-type stimulants; Earthworms; Toxicity.

6.1 Introduction MAP is an amphetamine type of psychoactive compound, which acts on the central nervous system as a stimulant (Pal et al., 2013; UNODC, 2013). The consumption and seizure of MAP has increased dramatically worldwide and in Australia over the past decade. Globally, MAP is the most commonly abused synthetic illicit compound after cannabis. The abuse of MAP is closely linked to serious human health, social, financial and security issues (UNODC, 2013). In recent years, the occurrence of MAP has been reported in several countries’ different environmental systems. The primary source of MAP contamination of the environment is believed to be human consumption and manufacturers’ illegal disposal of waste from their facilities (Pal et al., 2013). In most cases, illegal MAP manufacturers have been using groups of chemicals to produce substances, which may also be toxic and hazardous. These include

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solvents, precursors ephedrine and pseudoephedrine, metals and salts, acids and bases. A Consequently the illegal MAP production may generate considerable amounts of chemical wastes, which eventually end up in the environment and endanger the environmental systems. Normally, 3 kg of waste is produced when manufacturing 1 kg of MAP, depending on the producer’s skills (Pal et al., 2012). In the past, studies have concentrated on MAP toxicity, metabolism and its effects on human beings and animals. Typically, MAP administration to human beings and animals produces various peripheral and central effects, such as psychomotor activation, euphoria, decreased appetite, and hyperthermia (Quinton and Yamamoto, 2006). Human and animal chronic exposure to MAP has resulted in long-term damage to the dopaminergic and serotonergic systems, thus severely compromising nerve enzyme activity, monoamine content, and monoamine transporters function (Quinton and Yamamoto, 2006). Furthermore, acute MAP overdose and chronic administration to model species can cause mitochondrial dysfunction and cell death (Davidson et al., 2001). Although the toxicity of MAP to human beings and animals is well documented, data concerning its impact on the environmental system is still missing. The earthworm (E. fetida) was chosen for this experiment to evaluate the environmental effects of MAP because of this particular creature’s ecological significance, ease of use and availability (Yasmin and D’Souza, 2007). Additionally, earthworms are a regularly studied invertebrate with regard to uptake of soil contaminants because they can process large quantities of soil (Das et al., 2013). Generally, MAP manufacturers are burying the waste materials in soil or disposing of it into domestic waste disposal areas, which poses a serious threat to the environment because methamphetamine is persistent in soil and related soil processes (Pal et al., 2012; Pal et al., 2013). Since earthworms live in close contact with soil this could reveal the toxic effects of MAP on the natural environment. Despite lower concentrations of MAP being reported from the environmental system, its existence as a mixture of pollutants coupled with long-term contact may cause critical ill effects to microbes and pose a serious risk to human beings (Binelli et al., 2012). Thus, in this experiment we aim to evaluate the toxic effects and DNA damage in earthworms caused by MAP in soil.

6.2 Materials and methods 6.2.1 Chemicals and reagents Refer chapter 3.2.1. Comet assay kit was purchased from Trevigen, Inc. (Gaithersburg, MD, USA).

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6.2.2 Test soil The experimental soil was collected from Charlick Road, Crafers West (CW), South Australia. Refer chapter 4.2.2. The soils physical and chemical properties are presented in Table 4.1.

6.2.3 Test organism Refer chapter 5.2.2.

6.2.4 Toxicity assay Refer chapter 5.2.5. MAP toxicity to E. fetida was studied according to the OECD guidelines 1984 (OECD, 1984). MAP concentration tested in this experiment ranged between 5 and 200 mg kg-1 of soil.

6.2.5 Bioaccumulation test The toxicity and bioaccumulation of MAP to E. fetida was tested simultaneously based on an earlier documented procedure (Snyder et al., 2011). Worms were removed from test soil, washed and depurated (24 - 48 h), after 28 days of exposure. The depurated worms were weighed and frozen at -20 °C prior to extraction. The defrosted worms were extracted with 10 ml of chloroform: acetonitrile: methanol: acetic acid (80:10:9:1) in 40 ml glass vials. The extraction was carried out in the following sequence: the sample was homogenized, vortexed (30 sec), shaken in an orbital shaker (1 h) and then sonicated for 15 minutes at 30 °C. The vials were centrifuged and the aliquots were filtered through 0.22 µm Teflon filters. The supernatants were dried under nitrogen stream and dried extracts were reconstituted in 1 ml of HPLC grade of methanol for HPLC-MS analysis.

6.2.6 Reproduction test The reproduction test was conducted according to OECD guidelines 2004 (OECD, 2004). Soil moisture was maintained and animals were fed once a week with 2 g of oat flakes. Adult worms were removed from the test vessels after four weeks of soil exposure and containers were maintained in the state described above for four more weeks. At the end of the eighth week, the number of worm cocoons and juveniles was recorded.

6.2.7 Morphological and behavioural test

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Adult worms’ morphological changes and symptoms such as skin colour, texture, clitellum appearance, mucous secretion, body swelling, cuticle damage and fragmentation were monitored and recorded after eight weeks of soil exposure. E. fetida behavioural features such as ability to burrow, feeding habits, avoidance and movements of adult worms in the treated and control soils were also observed and documented throughout the test period.

6.2.8 Comet assay Refer chapter 5.2.9

6.2.9 Chemical analysis Refer chapter 4.2.4.

6.3 Results and discussion Toxicity data of MAP on microorganisms is scarce and to be specific there is no previous report on its effect on earthworms. However, a few reports on MAP toxicity to other aquatic organisms have indicated significant variances in toxicity to different organisms. As a consequence, it is difficult to compare this experimental outcome with previous relevant toxicity reports of MAP on earthworms. Bio-accumulation of MAP to E. fetida over a 28 days period at different concentrations is presented in Table 6.1. The bio-accumulation of MAP was increased with the increase in the concentration of MAP. At the highest test concentration (200 mg kg-1) MAP accumulation was 10.3 mg kg-1 followed by 4.1 and 2.0 mg kg-1 in 100 and 50 mg kg-1 exposed worms. The bioaccumulation of MAP in worm tissues at lower concentrations (5, 10, and 20 mg kg-1) ranged from as low as 0.2 to 0.9 mg kg-1. Bioaccumulation of MAP has been previously reported in arthropods, animals and human tissue (Gagliano-Candela and Aventaggiato, 2001; Moore et al., 1996) and may transfer into secondary level of bioaccumulation by means of entering into the food chain (Introna et al., 2001). Although the MAP accumulation in the eartworms of this study was low, the toxic nature of MAP along with its persistent contact to microorganisms is detrimental (Binelli et al., 2012). In this study, measured soil concentration of MAP was taken into account for all the calculations. There was no appreciable change (less than 10%) between nominal and measured soil MAP concentrations. MAP remained stable during the test periods, i.e. between 0 and 28 days (Table 6.1). The result is in agreement with Pal et al (2011) who reported that MAP persists in soil for a longer period depending on the soil type and condition. Percent weight changes of E. fetida in MAP treated soil over 4 weeks are presented in Figure 6.1. The percentage of weight loss was increased with increase in MAP concentration, while the control soil recorded a 1.3% increase in weight. At lower concentrations (5 - 20 mg 104

kg-1) the results ranged between 1.3 and 14.3%, while at 200 mg kg-1 36.3% of weight loss was recorded followed by 25.7 and 27.7% at 50 and 100 mg kg-1, respectively. The lethal concentration (LC50) value for MAP could not be calculated since no mortality was recorded throughout the experimental test concentrations. The similar type of weight loss effect was reported in earthworms exposed to pesticides and metals (Espinoza-Navarro and Bustos- Obregon, 2005; Klok et al., 2006). In nature, toxic properties of a chemical may cause death or trigger physiological stress that leads to significant weight loss (Frampton et al., 2006). In the present study the worms in the control recorded an increase in body weight which agrees with a previous report (Helling et al., 2000). Yasmin and D'Souza (2010) indicated changes in earthworms body weight is a sensitive tool to measure chemical toxicity rather than mortality. In this study, while no mortality was recorded the earthworm weight changes could indicate MAP’s toxic effects on earthworms. It has been reported that MAP exposure in human beings has resulted in substantial weight loss, sleeping disorders and loss of appetite (Brecht et al., 2004; Nordahl et al., 2003). These changes and behavioural variations cannot be ignored in MAP exposed earthworms. MAP has also been documented as initiating depletion of monoamines, neurotoxicity, oxidative stress and impairments in cognitive functions in human beings [Fleckenstein et al., 2007; Nordahl et al., 2003]. MAP may also have contributed to earthworms’ weight loss in this study. According to Bustos-Obregón and Goicochea (2002), this weight loss occurred due to toxic effects of the compound-altered muscular function which could reduce movements and feeding ability. A similar report in MAP exposed planarian species (Dugesia dorotocephala) where MAP inhibited locomotor movements and generation of energy by affecting dopamine and serotonin receptor sites (Janseen, 2008). The true effects of chemicals on earthworms can be assessed employing a reproduction test since species preservation and ecological balance requires a reproduction process (Landrum et al., 2006). MAP influence on earthworms’ reproduction is shown in Figure 6.2. The production of earthworm cocoons and juveniles declined with an increase in concentration of MAP. Overall the production of juveniles recorded a higher decline than cocoons due to MAP exposure. MAP up to 20 mg kg -1 did not have much impact on cocoons compared to juveniles. MAP at 50 mg kg -1 reduced 73.5% of cocoons and 88.9% juveniles, respectively, while the highest concentration (200 mg kg-1) reduced 100% of the juvenile’s production and 95.7% of cocoons production. Although the adult earthworms indicated no mortality, the reduction of cocoons and juveniles confirmed the subsidiary toxic effects of MAP exposure with the current results agrees well with previous reports on the effects of different chemicals on earthworms’ reproduction (Yasmin and D’Souza, 2007; Yasmin and D’Souza, 2010). The juvenile stage of earthworms is more susceptible to MAP than cocoons and adults,

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according to the above mentioned effects of organophosphorus (diazinon and chlorpyrifos) and cypermethrin insecticides (Booth and O’halloran, 2001; Zhou et al., 2008). Nonetheless the exact mechanism of how MAP acts on earthworms’ reproduction is still unknown. In this experiment, MAP exposed worms did exhibit significant morphological and behavioural change. It may be speculated that worms’ partner finding, mating posture, sperm production and quality may have led to the decrease in reproduction. Pesticide exposed worms have displayed parallel behavioural variations (Yasmin and D’Souza, 2010). In vertebrate species MAP primarily acts on the dopamine and serotonin transporter systems (Meredith et al., 2005), since they are important for normal functioning such as neuron message transfer, motor control, and physiological and behavioural control (Meredith et al., 2005; Rudnick, 2006). The results of this study on earthworms’ decreasing reproduction may be explained by the effects of MAP as described previously. However, more comprehensive analysis of the mechanism is required. Earthworms’ morphological and behavioural changes recorded for 28 days of MAP exposure in soil are shown in Table 6.1 and Figure 6.3. MAP up to 10 mg kg-1 did not significantly affect the earthworms’ morphological and behavioural traits. However, they showed marked morphological and behavioural changes when exposed to 20 mg kg-1 of MAP (Figure 6.3b). Higher MAP concentrations (>50mg kg-1) exhibited comparable effects on earthworms such as poor burrowing, coiling, skin dehydration, cuticle damage and fragmentation (Figure 6.3c and 3d). Similar effects were noted on earthworms chronically and acutely exposed to azodrin and chlorpyrifos (Rao and Kavitha, 2004; Rao et al., 2003). Several reports show that MAP when administered to animals and human beings causes psychosis, poor memory, violent behaviour, impulsiveness, oxidative stress, and impairments in cognitive function (Frenzilli et al., 2007; Nordahl et al., 2003). Furthermore, chronic usage of MAP can cause pathological cerebrovascular changes and haemorrhages in animals (Meredith et al., 2005), and skin diseases in human beings (Brecht et al., 2004). Apoptosis has also been noticed in MAP administered cell culture (Zhu et al., 2006). The present study on earthworms’ morphological and behavioural changes may be linked to the toxic effects of MAP, but the specific action by which such changes occurred require further investigation.

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Table 6.1 MAP soil concentration, bioaccumulation and biological parameters of earthworms. Soil Earthworm Nominal Measured concentration of concentration MAP of MAP mg kg-1 mg kg-1 0 day 28 days MAP Burrowing Coiling Skin Cuticle damage bioaccumulation dehydration and mg kg-1 fragmentation Control 0.0 ± 0.0 0.0 ± 0.0 0.0 ± 0.0 √ √ √ √ 5 4.5 ± 0.8 4.3 ± 0.9 0.2 ± 0.0 √ √ √ √ 10 8.9 ± 1.1 8.8 ± 0.7 0.3 ± 0.1 √ √ √ √ 20 18.8 ± 0.9 18.3 ± 0.4 0.9 ± 0.0 × × × √ 50 45.1 ± 1.7 43.8 ± 1.6 2.0 ± 1.2 × × × × 100 89.9 ± 1.6 89.3 ± 0.8 4.1 ± 0.5 × × × × 200 194.5 ± 1.8 192.0 ± 0.7 10.3 ± 1.1 × × × ×

Note: √ = no observed negative effect, × = negative effect

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Figure 6.1 MAP effect on earthworm weight changes (%) over control.

120.0

100.0

80.0

60.0

40.0

Weight changes (%) 20.0

0.0 Control 5 10 20 50 100 200 MAP concentration (mg kg-1)

Figure 6.2 MAP effects on earthworm reproduction (%) over control.

120.0

100.0

80.0 Cocoons Juveniles 60.0

40.0 Reproduction (%) 20.0

0.0 Control 5 10 20 50 100 200 MAP concentration (mg kg-1)

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Figure 6.3 Effects of MAP on earthworms following 28 days exposure in soil. (a) Control. (b) Earthworm coiling (20 mg kg-1). (c) Dehydrated earthworm (50 mg kg-1). (d) Cuticle damage and fragmentation (100 mg kg-1).

The results on earthworm’s DNA damage due to MAP over a 28-day soil exposure are presented in Figure 6.4. MAP influence on DNA damage in earthworms’ coelomic fluids was evaluated. The results showed that MAP was genotoxic to earthworms even at low concentration of 5 mg kg-1 and an equal level of olive tail movements was recorded in both 5 and 20 mg kg -1 of MAP concentration. A significant difference was noticed in the DNA of MAP exposed earthworms (p < 0.05) compared to the control. Figure 6.5a and 6.5b illustrates the minimum damage caused to earthworm’s coelomic fluid cells in the control sample, and MAP exposed earthworms’ DNA migration and damage as a result of strand breakage. The percentages of DNA damage and olive tail movements were more noticeable at 100 mg kg-1 than in the control. Our results agree well with previous studies reporting DNA damage to vertebrates after acute and chronic exposure to MAP (Tokunaga et al., 2008). MAP is toxic to rats’ multiple organs and various regions of brain cells even at single lower dosage administration by elevating lipid oxidation products (Gluck et al., 2001) and embryonic oxidative DNA damages were recorded in adult mice (Jeng et al., 2005). The present study finds that genetic damage in earthworms’ fluid cells is possibly the result of oxidative DNA damage of MAP. Generally, before these oxidative effects occur, MAP mainly acts on dopamine where auto-oxidation produces various toxic radicals such as peroxide, hydroxyl, superoxide and nitrogen (Thrash et al., 2009). These radicals are highly reactive with sugar, protein and lipids, resulting in many internal modifications. The major documented changes in animal systems are cell death, mitochondrial malfunction, metabolic energy reduction, and apoptosis (Imam et al., 2001; Riddle et al., 2006). In addition, the MAP oxidative process as it affects organisms inhibits mitochondrial ATP production; consequently very serious cell injury might occur (Thrash et al., 2009). This finding suggests that MAP induces cell damage in earthworms by affecting target sites on nuclei and mitochondria.

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Figure 6.4 DNA damage induced by MAP in earthworms. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA - ANOVA, p< 0.05 & olive tail movement - Dunett test, p < 0.05).

12.0 3.0 % tail DNA

10.0 * 2.5 Olive tail movement * d * 8.0 2.0

b c 6.0 1.5

% tail DNA 4.0 1.0

Olive tail movementOlive 2.0 0.5 a 0.0 0.0 control 5 20 100 MAP concentration (mg kg-1)

Figure 6.5 DNA damage in MAP exposed earthworms (E. fetida) as analysed by the comet assay. (a). Control with no or minimal DNA migrating into the tail region. (b). MAP 5 mg kg- 1exposed worms DNA migrating into the tail region as a result of strand breakage.

(a) (b)

6.4 Conclusion The present study revealed that MAP is toxic to E. fetida even at sub-lethal concentrations. Results have shown that there is a definite impact on earthworms’ life history parameters and cell level damage: weight changes, cocoon and juvenile production rate, and

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DNA damage. Furthermore, earthworms’ chronic exposure to MAP produced significant morphological and behavioural changes. In particular, from 20 mg kg-1 exposed worms exhibited poor burrowing, coiling and skin dehydration, and cuticle damages occurred from 100 mg kg-1. It can be concluded from these results that MAP is toxic to soil microbes and can enter the food chain when it accumulates in earthworms’ tissue.

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6.5 References

Binelli A, Pedriali A, Riva C, Parolini M. Illicit drugs as new environmental pollutants: Cyto- genotoxic effects of cocaine on the biological model Dreissena polymorpha. Chemosphere 2012; 86:906-11. Booth LH, O'halloran K. A comparison of biomarker responses in the earthworm Aporrectodeaca liginosa to the organophosphorus insecticides diazinon and chlorpyrifos. Environ Toxicol Chem 2001; 20:2494-2502. Brecht ML, O'Brien A, Von Mayrhauser C, Anglin MD. Methamphetamine use behaviours and gender differences. Addict Beha 2004; 29:89-106. Bustos-Obregón E, Goicochea RI. Pesticide soil contamination mainly affects earthworm male reproductive parameters. Asian J Androl 2002; 4:195-200. Das P, Megharaj M, Naidu R. Perfluorooctane sulfonate release pattern from soils of fire training areas in Australia and its bioaccumulation potential in the earthworm Eisenia fetida. Environ Sci Pollut Res 2013; 1-9. Davidson C, Gow AJ, Lee TH, Ellinwood EH. Methamphetamine neurotoxicity: necrotic and apoptotic mechanisms and relevance to human abuse and treatment. Bra Res Rev 2001; 36:1-22. Espinoza‐Navarro O, Bustos‐Obregón E. Effect of malathion on the male reproductive organs of earthworms, Eisenia foetida. Asian J Androl 2005; 7:97-101. Fleckenstein AE, Volz TJ, Riddle EL, Gibb JW, Hanson GR. New insights into the mechanism of action of amphetamines. Annu Rev Pharmacol Toxicol 2007; 47:681-698. Frampton GK, Jänsch S, Scott‐Fordsmand JJ, Römbke J, Van Den Brink PJ. Effects of pesticides on soil invertebrates in laboratory studies: a review and analysis using species sensitivity distributions. Environ Toxicol Chem 2006; 25:2480-2489. Frenzilli G, Ferrucci M, Giorgi FS, Blandini F, Nigro M, Ruggieri S, Murri L, Paparelli A, Fornai F. DNA fragmentation and oxidative stress in the hippocampal formation: a bridge between 3,4-methylenedioxymethamphetamine (ecstasy) intake and long-lasting behavioural alterations. Beha Pharmacol 2007; 8:471-481. Gagliano-Candela R, Aventaggiato L. The detection of toxic substances in entomological specimens. Int J Legal Med 2001; 114:197-203. Gluck MR, Moy LY, Jayatilleke E, Hogan KA, Manzino L, Sonsalla PK. Parallel increases in lipid and protein oxidative markers in several mouse brain regions after methamphetamine treatment. J Neurochem 2001; 79:152-160.

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Helling B, Reinecke S, Reinecke A. Effects of the Fungicide Copper Oxychloride on the Growth and Reproduction of Eisenia fetida (Oligochaeta). Ecotoxicol Environ Saf 2000; 46:108-116. Imam SZ, El‐Yazal J, Newport GD, Itzhak Y, Cadet JL, Slikker W, Ali SF. Methamphetamine‐ Induced Dopaminergic Neurotoxicity: Role of Peroxynitrite and Neuroprotective Role of Antioxidants and Peroxynitrite Decomposition Catalysts. Ann NY Acad Sci 2001; 939: 366-380. Introna F, Campobasso CP, Goff ML. Entomotoxicology. Forensic Sci Int 2001; 120: 42-47. Janssen S. Characterization of the effects of amphetamines in the planarian species Dugesia dorotocephala. Phd thesis 2008, University of Florida, USA. Jeng W, Wong AW, Ting-A-Kee R, Wells PG. Methamphetamine-enhanced embryonic oxidative DNA damage and neurodevelopmental deficits. Free Radic Biol Med 2005; 39: 317-326. Klok C, Van Der Hout A, Bodt J. Population growth and development of the earthworm Lumbricusrubellus in a polluted field soil: Possible consequences for the godwit (Limosalimosa). Environ Toxicol Chem 2006; 25:213-219. Landrum M, Cañas JE, Coimbatore G, Cobb GP, Jackson WA, Zhang B, Anderson TA. Effects of perchlorate on earthworm (Eisenia fetida) survival and reproductive success. Sci Total Environ 2006; 363:237-244. Meredith CW, Jaffe C, Ang-Lee K, Saxon AJ. Implications of chronic methamphetamine use: a literature review. Harvard Rev Psychiatry 2005; 13:141-154. Moore KA, Mozayani A, Fierro MF, Poklis A. Distribution of 3, 4- methylenedioxymethamphetamine (MDMA) and 3, 4-methylenedioxyamphetamine (MDA) stereoisomers in a fatal poisoning. Forensic Sci Int 1996; 83:111-119. Nordahl TE, Salo R, Leamon M. Neuropsychological effects of chronic methamphetamine use on neurotransmitters and cognition: a review. The J Neuropsychiatry Clinical Neurosci 2003; 15:317-325. Organisation for Economic Co-operation and Development, Earthworm acute toxicity test, OECD Guideline 207, 1984, Paris, France. Organisation for Economic Co-operation and Development, Earthworm reproduction test (Eisenia fetida), OECD Guideline 220, 2004, Paris, France. Pal R, Megharaj M, Kirkbride KP, Naidu R. Fate of 1-(1′, 4′-cyclohexadienyl)-2- methylaminopropane (CMP) in soil: Route-specific by-product in the clandestine manufacture of methamphetamine. Sci Total Environ 2012; 416:394-399. Pal R, Megharaj M, Kirkbride KP, Naidu R. Illicit drugs and the environment- A review. Sci Total Environ 2013; 463-464:1079-1092.

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Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor, and by-products in soil. Chemosphere 2011; 85:002-1009. Quinton MS, Yamamoto BK. Causes and consequences of methamphetamine and MDMA toxicity. The AAPS j 2006; 8:E337-E337. Rao JV, Kavitha PV. Toxicity of azodrin on the morphology and acetylcholinesterase activity of the earthworm (Eisenia foetida). Environ Res 2004; 96:323-327. Rao JV, Pavan YS, Madhavendra S. Toxic effects of chlorpyrifos on morphology and acetylcholinesterase activity in the earthworm, Eisenia foetida. Ecotoxicol Environ Safety 2003; 54:296-301. Riddle EL, Fleckenstein AE, Hanson GR. Mechanisms of methamphetamine-induced dopaminergic neurotoxicity. The AAPS J 2006; 8:E413-E418. Rudnick G. Serotonin transporters–structure and function. The J Membrane boil 2006; 213: 101-110. Snyder EH, O’connor GA, Mcavoy DC. Toxicity and bioaccumulation of biosolids-borne triclocarban (TCC) in terrestrial organisms. Chemosphere 2011; 82:460-467. Thrash B, Thiruchelvan K, Ahuja M, Suppiramaniam V, Dhanasekaran M. Methamphetamine- induced neurotoxicity: the road to Parkinson’s disease. Pharmacol Rep 2009; 61:966- 77. Tokunaga I, Ishigami A, Kubo SI, Gotohda T, Kitamura O. The peroxidative DNA damage and apoptosis in methamphetamine-treated rat brain. J Invest Med 2008; 55; 241-5. UNODC (United Nations Office on Drugs and Crime). World Drug Report 2013. http://www.unodc.org/unodc/secured/wdr/wdr2013/World_Drug_Report_2013.pdf. [Accessed on October 25, 2013]. Yasmin S, D’Souza D. Effect of pesticides on the reproductive output of Eisenia fetida. Bull Environ Contam Toxicol 2007; 79:529-532. Yasmin S, D'souza D. Effects of pesticides on the growth and reproduction of earthworm: a review. App Environ Soil Sci 2010:1-9. doi:10.1155/2010/678360. Zhou S, Duan C, Wang X, Michelle WHG, Yu Z, Fu H. Assessing cypermethrin-contaminated soil with three different earthworm test methods. J Environ Sci 2008; 20:1381-1385. Zhu ZP, Xu W, Angulo JA. Methamphetamine-induced cell death: selective vulnerability in neuronal subpopulations of the striatum in mice. Neuroscience 2006; 140:607-622.

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CHAPTER 7

Earthworm chronic exposure to Pseudoephedrine (PSE) – impacts on life parameters and DNA

Abstract Pseudoephedrine (PSE) is an over-the-counter decongestant drug, commonly used for asthma, nasal congestion and obesity. It is well-known that PSE is used in illegal methamphetamine synthesis and it produces psychostimulant effects if taken in overdoses. To date there have been no reports on the environmental toxicity of PSE. This study examines the chronic effects of PSE exposure to earthworms. Earthworms were exposed to PSE treated soil at 5 to 200 mg kg-1. The results showed at higher concentrations, earthworm mortality was absent. However, the higher test concentrations (100 and 200 mg kg-1) weight loss ranging between 5 - 9.6% was recorded, and also morphological and behavioural changes were observed, such as coiling, skin dehydration and discolouration. Reproductive potential was also declined especially above 50 mg kg-1 concentration. The exposure at higher concentrations of 50 - 200 mg kg-1 considerably reduced both cocoon production and juvenile emergence by 14.3 - 17.9% and 27.2 - 33.8%, respectively, suggesting PSE was more toxic to juveniles. PSE accumulation in worm tissue was also noted at higher concentrations but significant damages to earthworms’ DNA was recorded even at 5 mg kg-1 concentration. Results suggest that pseudoephedrine exposure results in toxicity to earthworms.

Keywords: Illicit drugs; Pseudoephidrine; DNA; Earthworms; Toxicity.

7.1 Introduction Pseudoephedrine (PSE) is a sympathomimetic compound mainly found in cold-flu and nasal decongestant over-the-counter medicines (Moser and Rayburn, 2007). In recent years PSE has been used as a preferred precursor for the illegal manufacture of methamphetamine(MAP). MAP production and its use is steadily increasing globally and in Australia (UNODC, 2013). This has generated a direct demand for PSE for illegal markets. Internationally, India and Germany are contributing 90% of the PSE export followed by Taiwan and China (INCSR, 2012). Only the smallest portions of exported PSE have been employed in the production of legal medicines. Although PSE is a legitimate medicine, excessive and continuous exposure may be harmful to human beings, animals and soil microbes.

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Clandestine waste materials containing PSE are usually buried in soil and in recent years the occurrence of PSE in the natural environment has also been documented (Pal et al., 2013). To date, while no report on the ecological toxicity of PSE has been published, although several reports exist on toxic effects to humans and animals. PSE use in human beings has produced various adverse interactions, ranging from mild dermatitis, headache, and dizziness to serious cardiovascular and neurological problems, epidermal necrolysis, hepatitis, myocardial infarction, chest pain, stroke and death (Fukushima, 2004; Ben Salem et al., 2008). Hypertension was the most frequently recorded adverse reaction followed by tachycardia, myocardial infarction, stroke and seizure (Fukushima, 2004; Andraws et al., 2005). In animals, PSE administration causes non- convulsive seizure, CNS oxygen toxicity, reduction in sperm quality and quantity, increased apoptotic activity and developmental toxicity (Nudmamud-Thanoi and Thanoi, 2012; Pilla et al., 2013). The literature on PSE’s ill effects to the environment is void and for this reason it is worth exploring the environmental fate and behaviour of PSE using ecologically-related organisms. To assess the ecotoxicity of PSE, the earthworm (E. fetida) was selected because it lives in close contact with soil and through dermal or gut, processes large quantities of soil. These ecologically relevant characteristics of earthworms enable us to investigate soil contaminants like PSE’s bioaccumulation, its impacts on life parameters and damage to cells. Furthermore, E. fetida is very common in agricultural and garden soil, easy to handle and culture in laboratory conditions. Since most of the illicit drug wastes are reaching the soil in different ways, the earthworm may be the ideal organism to test the environmental impacts of PSE. Despite the fact that PSE degrades relatively faster in soil (Pal et al., 2011), its adverse toxic effects on earthworms are yet to be understood. Thus, the present study aims to investigate PSE’s toxic effects on E. fetida life parameters and DNA.

7.2 Materials and methods

7.2.1 Soil collection and preparation The test soil was collected from Charlick Road, Crafers West (CW), South Australia. Refer chapter 4.2.2. The soils physical and chemical properties are given in Table 4.1.

7.2.2 Reagents Refer chapter 3.2.1

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7.2.3 Earthworms Refer chapter 5.2.2

7.2.4 Toxicity and bioaccumulation test Refer chapter 6.2.4 and 6.2.5

7.2.5 Reproduction test Refer chapter 6.2.6

7.2.6 Morphological and behavioural test Refer chapter 6.2.7

7.3 Results and discussion The weight changes in percentage terms of E. fetida over the 4-week period using different concentrations of PSE are presented in Figure 7.1. Weight loss was observed in higher concentrations of 50, 100 and 200 mg kg-1 that ranged between 1.1 - 9.6%, while in the control sample and in the lower concentrations of 5, 10 and 20 mg kg -1 exposed worms did indicate minor weight gain. All earthworms were found to be alive throughout the experimental duration for all concentrations tested. The present results suggest that earthworm weight loss in higher concentrations was probably due to the toxic influence of PSE. In human beings, PSE containing dietary supplements induced weight loss ranging between 0.6 to 0.9 kg/month (Dyck, 2000). This weight loss is believed to have occurred due to an increased metabolic rate, stimulatory effects, thermogenesis and PSE’s ability to suppress the appetite (Astrup et al., 1992). A similar toxic outcome has been reported in earthworms exposed to pesticides and metals (Espinoza‐Navarro and Bustos‐Obregón, 2005). Measurement of earthworms’ weight loss is a better indicator of a chemical’s toxic effects than mortality (Yasmin and D'Souza, 2010) and it is a valuable sign of physiological strain exerted by the chemical based on its toxicity level and exposure period (Frampton et al., 2006). The physiological stress of chemicals on earthworms may alter muscular function and feeding ability (Bustos-Obregón and Goicochea, 2002) and consequently may result in weight loss. Earthworms’ increase in weight in control is consistent with the findings of Helling et al. (2000). Bioaccumulation of PSE in E. fetida over the 4-week period at different concentrations is presented in Table 7.1. Lower test concentrations of PSE, i.e. 5 and 10 mg kg-1 did not record any accumulation in earthworms’ tissue, while 20 - 200 mg kg-1 showed little 117

accumulation ranging between 0.1 to 0.6 mg kg-1. Information is still lacking on PSE accumulation in microorganisms, animals and human beings. However, toxicity of PSE, a structurally and pharmacologically similar compound to amphetamine (Ruksee et al., 2008) has been reported in insects, animals and human beings (Gagliano-Candela and Aventaggiato, 2001). In general, substances accumulated in an organism always tend to transfer to the next biotic level in the food chain (Introna et al., 2001). Overall, in this study PSE demonstrated less or no accumulation (5 and 10 mg kg-1) in earthworms’ tissue, which may be explained by PSE’s short period of soil persistence and half-life (Pal et al., 2011). PSE’s chemical properties enable it to be quickly eliminated from animals and human beings (Ruksee et al., 2008). In humans, 81 - 96% of administered PSE is excreted unchanged within

24 h with a half-life (t1/2) of 4.2 h. (Chan et al., 2008). Moreover, PSE has similar pharmacological effects to amphetamines, so it is unsafe to ignore the consequences based on its accumulation in overdose exposure to earthworms.

Figure 7.1 PSE effects on earthworms weight changes (%) over control.

120.0

100.0

80.0

60.0

40.0

Weight changes (%) 20.0

0.0 Control 5 10 20 50 100 200 PSE concentration (mg kg-1)

The results of this experiment confirmed that higher test concentrations of PSE affected reproduction capacity, morphology and behaviour in earthworms (Table 7.1 and Figure 7.2). PSE up to 20 mg kg-1 did not affect the earthworms’ reproduction of either cocoons or juveniles. Test concentrations of 50,100 and 200 mg kg-1 significantly redued reproduction in earthworm. PSE test concentration of 50 - 200 mg kg-1 decreased earthworm cocoons and juveniles, respectively, in ranges between 14.3 - 17.9% and 27.2 - 33.8%. The juvenile stage was more susceptible to PSE compared to the cocoon stage. Furthermore, worms exposed to PSE at 100 and 200 mg kg-1 displayed significant morphological and behavioural changes such as skin dehydration, discolouration, coiling and poor burrowing (Table 7.1), while lower test concentrations of 5 - 50 mg kg -1 did not demonstrate any significant effects. The findings 118

in this study on earthworms’ life parameters may have been indirectly caused by an overdose of PSE with subsequent sympathomimetic outcomes.

PSE is a well-known adrenergic agonist that can enhance the release of norephedrine from the sympathetic neuron (Fitzgerald et al., 2006). Norephedrine and adrenergic receptors can reduce spermatogenesis and the steroidogenic capacity of Leydig cells (Mhaouty-Kodja et al., 2007; Nudmamud-Thanoi and Thanoi, 2012). In a previous study a higher dose (180 mg kg-1) of PSE administered to rats induced abnormal sperm morphology, reduction in sperm count and increased apoptotic activities (Nudmamud-Thanoi and Thanoi, 2012). Earthworms’ reproduction and behavioural tests are real and rapid indicators that can be used to assess the risk of chemical-contaminated soil (Landrum et al., 2006). The present study’s results agree well with previous analyses of pesticides’ dangerous impacts on earthworms’ reproduction, juveniles’ vulnerability, and morphological and behavioural changes (Rao and Kavitha, 2004; Yasmin and D’Souza, 2007). Yasmin and D'Souza (2010) stated that worms exposed to pesticides had reduced reproduction capacity, possibly due to these substances’ toxicity affecting partner finding, mating posture and sperm quality and quantity. It may be speculated that the above outcomes in this present study did trigger a decline in earthworms’ reproduction ability. Further, PSE is structurally and pharmacologically similar to amphetamines except that it demonstrates less potency (Langston, 2001). It is well established that amphetamines given to vertebrates causes stimulatory effects, suppresses appetite, increases psychosis and violent behaviour, weakens memory, elevates impulsivity, oxidative stress, and creates impairments in cognitive function (Easton and Marsden, 2006; Frenzilli et al., 2007). Chronic use of PSE has induced pathological cerebrovascular changes and signs of haemorrhage in animals (Meredith et al., 2005), and skin diseases in human beings (Brecht et al., 2004). These results may occur in higher concentrations of PSE that induce morphological and behavioural changes in earthworms. Earlier studies did report that overdoses of PSE in vertebrates induced intracranial haemorrhage, skin complication and behavioural changes (Fitzgerald et al., 2006; Ben Salem et al., 2008).

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Table 7.1 PSE soil concentration, bioaccumulation and biological parameters of earthworms

Soil Earthworm

PSE Measured concentrations of Nominal bioaccumulation Morphological and behavioural parameters PSE mg kg-1 -1 concentrations mg kg of PSE Skin -1 Skin Swelled mg kg 0 day 28 days Burrowing colour to texture clitellum yellowish

Control 0.0 ± 0.0 0.0 ± 0.0 0.00 ± 0.0 √ √ √ √

5 4.6 ± 0.4 0.3 ± 0.2 0.00 ± 0.0 √ √ √ √

10 8.1 ± 0.1 0.3 ± 0.8 0.00 ± 0.0 √ √ √ √

20 17.8 ± 0.7 0.8 ± 0.9 0.15 ± 0.0 √ √ √ √

50 44.1 ± 1.7 1.6 ± 0.9 0.57 ± 0.3 √ √ √ √

100 92.9 ± 0.6 3.2 ± 1.4 0.66 ± 0.1 √ × × √

200 191.5 ± 0.1 4.7 ± 1.1 0.69 ± 0.1 √ × × √

Note: √ = no observed negative effect, × = negative effect

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Figure 7.2 PSE effects on earthworm reproduction (%) over control

120.0 Cocoons 100.0 80.0 Juveniles 60.0 40.0

Reproduction (%) 20.0 0.0

PSE concentration (mg kg-1)

Chronic PSE exposure caused DNA damage in earthworms (Figure 7.3). Results showed that PSE was genotoxic to earthworms even at a low concentration of 5 mg kg-1 and the olive tail movements increased when the concentration also increased. A significant difference was observed in the DNA of PSE exposed worms (p < 0.05) compared to the control. Figure 7.4a & b show only minimal damage to earthworms’ coelomic fluid cells in the negative control and DNA migration and damage as a result of strand breakage in PSE exposed worms (positive control). The percentages of DNA damage and olive tail movements were more noticeable at 100 mg kg-1 compared to the control. Despite the fact there is no literature on genotoxicity of PSE to various test-systems, a few studies have reported that PSE induces apoptosis activities, renal tubular cell degeneration, and cytotoxicity to human neuroblastoma SH-SY5Y and rat myoblastoma H9c2 (2-1) cell lines (Fukushima, 2004; Soni et al., 2004; Nudmamud- Thanoi and Thanoi, 2012). The apoptosis and cytotoxicity activities are indicators of cell death, necrosis and signals of DNA strand breakage or fragmentation (Brown and Attardi, 2005).

Given the evidence in this study, PSE can induce cell damage in earthworms’ coelomic fluid cells, possibly as a result of oxidative DNA damage (Gardner et al., 1997). It has also been reported to be a main precursor for methamphetamine manufacturing which produces very similar effects in human beings and animals (Zhang et al., 2008). Several studies have documented that methamphetamine can induce DNA damage through oxidative effects (Jeng et al., 2005; Tokunaga et al., 2008). However, ephedrine can have various side-effects ranging from minor to fatal, but short-term exposure does not induce genetic damage (Soni et al., 2004; Radaković et al., 2011). The present findings must be interpreted cautiously because the comet assay result does not necessarily predict the mutagenic effects of PSE. Nonetheless, it can be stated that PSE can damage earthworms’ coelomic fluid cells. 121

Figure 7.3 DNA damage induced by PSE in earthworms. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA - ANOVA, p < 0.05 & olive tail movement - Dunett test, p < 0.05)

Figure 7.4 DNA damage in PSE exposed earthworms (E. fetida) as analysed by the comet assay. (a). Control with no or minimal DNA migrating into the tail region. (b). PSE exposed (5 mg kg-1) worms DNA migrating into the tail region as a result of strand breakage.

(a) (b)

7.4 Conclusion This study is the first to describe the chronic toxicity effects of pseudoephedrine on an earthworm life parameters and DNA. Overdoses of pseudoephedrine can severely disrupt 122

their weight, reproduction, morphology and behaviour. Even at smaller concentrations of 5 mg kg -1) pseudoephedrine, its chronic exposure can cause significant damage to earthworms’ coelomic fluid cells. Finally, the results obtained in this investigation provide important and useful knowledge concerning the toxicological effects of pseudoephedrine, given that it is commonly used as a decongestant.

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7.5 References

Andraws R, Chawla P, Brown DL. Cardiovascular effects of ephedra alkaloids: a comprehensive review. Prog Cardiovasc Dis 2005; 47:217-225. Astrup A, Buemann B, Christensen NJ, Toubro S, Thorbek G, Victor OJ, Quaade F. The effect of ephedrine/caffeine mixture on energy expenditure and body composition in obese women. Metabolism 1992; 41:686-688. Ben Salem C, Slim R, Denguezli M, Sriha B, Hmouda H, Bouraoui K. Pseudoephedrine‐ induced acute generalized exanthematous pustulosis. Int J Dermatol 2008; 47:418- 419. Brecht ML, O'Brien A, Von Mayrhauser C, Anglin MD. Methamphetamine use behaviours and gender differences. Addict Behav 2004; 29:89-106. Brown JM, Attardi LD. The role of apoptosis in cancer development and treatment response. Nature Rev. Cancer 2005; 5:231-237. Bustos-Obregón E, Goicochea RI. Pesticide soil contamination mainly affects earthworm male reproductive parameters. Asian J Androl 2002; 4:195-200. Chan KH, Pan RN, Hsu, MC, Hsu, KF. Urinary elimination of ephedrines following administration of the traditional Chinese medicine preparation Kakkon-to. J Anal Toxicol 2008; 32:763-767. Dyck DJ. Dietary fat intake, supplements, and weight loss. Canadian J Appl Physiol 2000; 25: 495-523. Easton N, Marsden CA. Ecstasy: are animal data consistent between species and can theytranslate to humans? J Psychopharmacol 2006; 20:194-210. Espinoza‐Navarro O, Bustos‐Obregón E. Effect of malathion on the male reproductive organs of earthworms, Eisenia foetida. Asian J Androl 2005; 7:97-101. Fitzgerald KT, Bronstein AC, Flood AA. “Over-The-Counter” Drug Toxicities in Companion Animals. Clin Tech Small Ani Prac 2006; 21:215-226. Frampton GK, Jänsch S, Scott‐Fordsmand JJ, Römbke J, Van Den Brink PJ. Effects of pesticides on soil invertebrates in laboratory studies: a review and analysis using species sensitivity distributions. Environ Toxicol Chem 2006; 25:2480-2489. Frenzilli G, Ferrucci M, Giorgi FS, Blandini F, Nigro M, Ruggieri S, Murri L, Paparelli A, Fornai F. DNA fragmentation and oxidative stress in the hippocampal formation: a bridge between 3,4-methylenedioxymethamphetamine (ecstasy) intake and long-lasting behavioural alterations. Behav Pharmacol 2007; 8:471-481.

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Fukushima K. Bioactivity of Ephedra: Integrating Cytotoxicity Assessment with Real-Time Biosensing. Master of Science thesis (2004). Faculty of the Graduate School of the University of Maryland, College Park, MD, 20742. Gagliano-Candela R, Aventaggiato L. The detection of toxic substances in entomological specimens. Int. J. Legal Med 2001; 14:197-203. Gardner AM, Xu FH, Fady C, Jacoby FJ, Duffey DC, Tu Y, Lichtenstein A. Apoptotic vs. nonapoptotic cytotoxicity induced by hydrogen peroxide. Free Rad. Bio. Med 1997; 22: 73-83. Helling B, Reinecke S, Reinecke A. Effects of the Fungicide Copper Oxychloride on the Growth and Reproduction of Eisenia fetida (Oligochaeta). Ecotoxicol Environ Saf 2000; 46:108-116. INCSR (International Narcotics Control Strategy Report), 2012. Drug and chemical control. Volume 1. (http://www.state.gov/j/inl/rls/nrcrpt/2012/index.htm) [Accessed on 20th January 2014] Introna F, Campobasso CP, Goff ML. Entomotoxicology. Forensic Sci. Int 2001; 120: 42-47. Jeng W, Wong AW, Ting-A-Kee R, Wells PG. Methamphetamine-enhanced embryonic oxidative DNA damage and neurodevelopmental deficits. Free Rad Bio Med 2005; 39: 317-326. Landrum M, Cañas JE, Coimbatore G, Cobb GP, Jackson WA, Zhang B, Anderson TA. Effects of perchlorate on earthworm (Eisenia fetida) survival and reproductive success. Sci Total Environ 2006; 363:237-244. Langston C. Pharmacotherapy of neonates and pregnant animals. Proc 19th ACVIM Forum, Denver CO 2001; 80-82. Meredith CW, Jaffe C, Ang-Lee K, Saxon AJ. Implications of chronic methamphetamine use: a literature review. Harvard Rev Psychiatry 2005; 13:141-154. Mhaouty-Kodja S, Lozach A, Habert R, Tanneux M, Guigon C, Brailly-Tabard S, Maltier JP, Legrand-Maltier, C. 2007. Fertility and spermatogenesis are altered in α1b-adrenergic receptor knockout male mice. J Endocrinol 2007; 195:281–292. Moser B, Rayburn J. Evaluation of developmental toxicity of interaction between caffeine and pseudoephedrine using frog embryo teratogenesis assay-Xenopus (Fetax). Bioscience 2007; 78:1-9. Nudmamud-Thanoi S, Thanoi S. Pseudoephedrine induces sperm abnormalities, lower sperm counts and increased apoptosis in rat testis. Cell Tissue Res 2012; 349:625- 630. Pal R, Megharaj M, Kirkbride KP, Naidu R. Illicit drugs and the environment- A review. Sci Total Environ 2013; 463-464:1079-1092.

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Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor, and by-products in soil. Chemosphere 2011; 85:1002-1009. Pilla R, Held HE, Landon CS, Dean JB. High doses of pseudoephedrine hydrochloride accelerate onset of CNS oxygen toxicity seizures in unanesthetized rats. Neurosci 2013; 246:391-396. Radaković M, Đelić N, Stanimirović Z, Plećaš-Solarović B, Spremo-Potparević B, Živković L. & Bajić V. Evaluation of the effects of ephedrine on human lymphocytes in the comet assay. Acta Vet 2011; 61:363-371. Rao JV, Kavitha P. Toxicity of azodrin on the morphology and acetylcholinesterase activity of the earthworm (Eisenia foetida). Environ Res 2004; 96:323-327. Ruksee N, Tongjaroenbuangam W, Casalotti SO, Govitrapong P. Amphetamine and pseudoephedrine cross-tolerance measured by c-Fos protein expression in brains of chronically treated rats. BMC Neurosci 2008; 9:99. Soni MG, Carabin IG, Griffiths JC, Burdock G A. Safety of ephedra: lessons learned. Toxicol Lett 2004; 150:97-110. Tokunaga I, Ishigami A, Kubo SI, Gotohda T, Kitamura O. 2008. The peroxidative DNA damage and apoptosis in methamphetamine-treated rat brain. J Invest Med 2008; 55: 241-5. UNODC (United Nations Office on Drugs and Crime). World Drug Report 2013. United Nations Publication. http://www.unodc.org/unodc/secured/wdr/wdr2013/World_Drug_Report_2013.pdf. [Accessed on October 25, 2013]. Yasmin S, D’souza D. Effect of pesticides on the reproductive output of Eisenia fetida. Bull Environ Contam Toxicol 2007; 79:529-532. Yasmin S, D'souza D. Effects of pesticides on the growth and reproduction of earthworm: a review. Appl Environ Soil Sci 2010; 1-9. Zhang JX, Zhang DM, Han XG. Identification of impurities and statistical classification of methamphetamine hydrochloride drugs seized in China. Forensic Sci Int 2008; 182: 13-19.

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CHAPTER 8

Acute toxicity and genotoxicity of amphetamine-type stimulant methamphetamine and its precursor pseudoephedrine to Daphnia carinata

Abstract This work is the first to report on acute toxicity of methamphetamine (MAP) and pseudoephedrine (PSE) to a cladoceran, Daphnia carinata. The study was conducted both in cladoceran and natural water collected from local creeks. The toxicity of MAP and PSE followed the order: cladoceran culture medium ˃ sterile natural water ˃ non-sterile natural water. MAP and PSE were relatively less toxic in non-sterile compared to sterile natural water, which could be due to the influence of varied physico-chemical parameters of natural water. In all the test media, MAP and PSE were found to be stable. In terms of genotoxicity, MAP and PSE induced significant DNA damage and olive tail movement to D. carinata at 0.25 - 1.0 mg L-1 and 0.5 - 1.0 mg L-1 in water exposure as compared to controls. It is clear that even low level chronic exposure of these compounds to D. carinata cause serious harmful effects including genetic material damages.

Keywords: Methamphetamine; Pseudoephedrine; Illicit drugs; Daphnia carinata; Toxicity; Comet assay

8.1 Introduction Illicit drugs are those for which nonmedical use is prohibited by national and international law, and mainly belong to the classes of opiates, cocaine, cannabis, amphetamine-type substances (ATS) and ecstasy-group substances (Pal et al.,2011; Hall et al., 2008). Drug abuse causes innumerable social problems, such as rising health care costs, organized crime, local violence, road fatalities and economic damage (van Nuijs et al., 2011). The recent world drug report has estimated that between 153 to 300 million people (aged 15- 64) used illicit drugs at least once in 2012 (UNODC, 2013). Among these, cannabis was the most widely used illicit drug (120 -224 million people) followed by ATS, opioids (opium, heroin and prescription opioids) and cocaine (UNODC, 2013). In Australia, ATS remains the second most used illicit drug after cannabis and in recent years the number of clandestine laboratories producing these compounds has raised several-fold (ACC, 2011-12). 127

MAP (N-methyl-1-phenyl-propan-2-amine, or methylamphetamine, C10H15N) is the most widely consumed synthetic stimulant in the world and very often is synthesized from easily available precursors like ephedrine and pseudoephedrine (EMCDDA, 2009). PSE

((1S,2S)-2-methylamino-1 -phenylpropan-1-ol, C10H15NO), remains the most popular precursor chemical involved in the manufacture of MAP, because it is easy to procure as over-the-counter cold and flu medications (Vearrier et al., 2012), is easily converted into MAP and because it is converted into the more biologically-active S-enatiomer (Stojanovska et al., 2013). Manufacturing MAP in a clandestine laboratory is much easier which require minimal effort and is easy to hide, and minimal space to set up than the production of plant-based drugs such as heroin and cocaine (UNODC, 2008). In 2010, MAP seizures worldwide increased to 45 tonnes compared to 31 tonnes in 2009 (EMCDDA, 2012). Illicit drugs and their metabolites are the latest group of emerging pollutants identified in the water ecosystem demanding more attention (Boles and Wells, 2010). They enter the wastewater network as unaltered drugs and/or their active metabolites by human excretion after illegal consumption or by accidental or deliberate disposal from clandestine drug laboratories (Pal et al., 2013, Boles and Wells, 2010). Worldwide several studies have demonstrated the presence of drugs of abuse in wastewater, surface water, drinking water, sludge, and air al (Pal et al., 2013), which may pose potential ecotoxicological dangers to humans and wildlife from chronic low level exposure (Pomati et al., 2006). Worldwide MAP concentration in untreated wastewater ranges from 3 ng L-1 to 800 ng L-1 (Pal et al., 2013), and in Australian wastewater the highest recorded concentration was 4108 ng L-1 (Irvine et al., 2011). In soil, MAP persists for a longer period of time (Pal et al., 2011) and may have potential pharmacological effects when combined with similar other compounds and pharmaceuticals present in surface waters (Zuccato et al., 2008). Though several reports are available on the presence of illicit drugs in water bodies worldwide, reports on the ecotoxicity of illicit drugs are lacking and there is no systematic information on potential harmful effects on aquatic organisms (Binelli et al., 2012). To date only a few reports are available on the ecotoxicity of ATS, cocaine, and morphine on aquatic organisms. The fresh water flea, D. carinata, is a popular aquatic test organism. These fleas occur in rivers, creeks and fresh water lakes, and serve as sentinel organisms in the natural environment. Furthermore, there are a number of other reasons for using D. carinata as a test organism, for example, it is common in Australian freshwater, it is an indigenous species, it is somewhat larger than D. magna, it has a high rate of neonate production, and it is easy to handle (Phyu et al., 2004). The environmental concentration of illicit drugs could well rise in the future, but data are lacking on the fate and behaviour of these compounds in aquatic

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ecosystems. This present study was designed to critically investigate the acute and geno - toxicity of MAP and PSE to D. carinata and their stability in both cladoceran medium water as well as natural waters.

8.2 Materials and methods 8.2.1 Test organism and culture condition D. carinata was cultured in cladoceran water (Caceres et al., 2007). The daphnid cultures were maintained and experiments conducted with 16 h light: 8 h dark photoperiod at 22 ± 2 °C, in glass containers (Caceres et al., 2007). Bold’s basal medium cultured green algae Raphidocelis subcapitata (Meharaj et al., 2000) was used as feed and medium water was renewed three times a week. The D. carinata used for all experiments were neonates of less than 24 h old.

8.2.2 Test compounds Refer chapter 3.2.1

8.2.3 Test water The test waters were cladoceran culture medium and two natural waters collected from Morialta Falls (MF) and Kangaroo Creek (KC), Adelaide. Both natural waters were collected in 20 L acid-washed high-density polyethylene containers.

8.2.4 Acute toxicity test The test method described by Caceres et al. (2007) was employed in this study to investigate the acute toxicity of MAP and PSE on fresh water D. carinata. The study was conducted in concentrations ranging between 5 and 400 mg L -1 of MAP and PSE in water. Survival tests of daphnia were conducted by introducing neonates in a 40 mL glass vial containing 20 mL of test solution maintained at 22 ± 2 °C following 16 h light and 8 h dark cycle with 600 Lux light intensity. Experimental solution without chemical served as the controls. To reduce evaporation of the test solution, transparent plastic film was used to cover all test vessels. Ten organisms were introduced into each vial and the experiment was conducted in triplicate. Acute toxicity of MAP and PSE was decided by death of daphnids used as the end point. Daphnids showing no active movement within 15 seconds after mild stirring were considered as immobile (dead). The numbers of immobile daphnids were recorded for each treatment and control after 24 h and 48 h. The data collected from each treatment were

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analysed statistically to identify the concentration of test chemicals (LC50) that immobilised 50% of the daphnids.

8.2.5 Chemical stability and analysis of MAP and PSE The role of microorganisms in the stability of MAP and PSE was tested in medium water as well as filter-sterilised (passed through 0.45 µm sterile Millipore filter) and non- sterilised natural waters at concentration levels of 5 and 100 mg L-1, respectively. All the water samples (cladoceran water and two natural waters) were tested for heavy metals, dissolved organic carbon (DOC), major anions (nitrate, phosphate, sulphate, fluoride, chloride and bromide) after passing through 0.45 µm Millipore filters. Inductively coupled plasma mass spectrometer (Agilent 7500 series) and ion chromatography (ICS 2000 series Dionex Ion Chromatograph System, USA) were used to determine heavy metals and major anions present in the waters. The dissolved organic carbon in the sample was determined with a total organic carbon analyser (OI Analytical, USA). The samples were incubated under the same experimental conditions. Unspiked water served as the controls and the study was conducted in triplicate. After 48 h, the stability of test chemicals were checked using high performance liquid chromatography (HPLC) equipped with a mass spectrometer (MS). Refer chapter 4.2.4. The limits of quantitation (LOQ) were set 5 ng L-1 for PSE and MAP, after established the concentration of the analytes in distilled water that gives rise to peak height with an S/N of 10. The two most responsive transitions per compound were used for quantitation of analytes. In this method the analytes average recoveries were 75 ± 2.3 (PSE) to 89 ± 1.6 (MAP) when n = 3. The analytical method for PSE and MAP showed a good linearity (r2 > 0.99) and repeatability ranged between 1.5 to 5.4%.

8.2.6 Comet assay DNA damage and tail movement were evaluated by the use of alkaline single cell gel electrophoresis or comet assay, according to the method of Singh et al. (1988) and as adapted by Park et al. (2010) for Daphnia. To perform comet assay, the 24 h neonates were exposed to test chemicals at concentrations ranging from 5 and 1000 µg L-1 for 96 h. Tests were conducted in triplicate with ten organisms per replicate. Organisms were placed in 1 mL of phosphate-buffered saline (PBS) containing 20 mM EDTA and 10% dimethyl sulfoxide (DMSO), and disintegrated mechanically by grinding. The suspension was centrifuged at 10000 rpm for 30 seconds (Thermo Scientific Heraeus® Fresco™ 21) to separate the debris and cells. Fifty micro litres of the cells was mixed with 150 µL of low melting agarose and mixed thoroughly by pipetting. Comet assay slides were coated with 50 µL of cells-agarose

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suspension and allowed it to solidify at 4 °C for 5 mins. The alkaline comet assay was performed according to the manufacturer’s instructions (Trevigen comet assay protocol, 8405 Helgerman Ct.). At the end of the assay, the slides were analysed using a fluorescence microscope (Olympus BX41) at 10X magnification. DNA damage was expressed as the tail and olive tail movement using an image analysis computerized method by CometScore software (TriTek Corp., USA).

8.3 Results and discussion

8.3.1 Physico-chemical properties of water Physico-chemical properties of test waters are presented in Table 8.1. Water quality aspects in both natural waters (MF and KC) varied noticeably from cladoceran water. MF and KC natural water had 4.0 and 8.6 mg L-1 of dissolved organic carbon (DOC) whereas cladoceran water contained less than 0.38 mg L-1. The electrical conductivity (EC) in MF and KC natural waters were 10- and 6-times greater than cladoceran water, respectively. Correspondingly, magnesium, sulfate, sodium and chloride content in MF and KC natural waters were 12-19, 7-16, 5-9, and 3-8 times more than the cladoceran water, respectively. Fluoride, nitrite, nitrate, phosphate, zinc, and lead were not detected in natural or cladoceran waters. Natural waters MF and KC had varied aqueous constituents (DOC and suspended particulate matter) that may have reduced toxic response of the MAP and PSE by reducing bioavailability to D. carinata. Earlier studies demonstrated the marked effects of DOC and sediments on chemical toxicity reduction (Nikkila et al., 2001; Phyu et al., 2004). In addition, natural waters used in this study had slightly higher pH and high chloride along with DOC, which may decrease the chemical’s toxicity. Similar effects have been reported previously for silver toxicity to rainbow trout (Oncorhynchus mykiss), fathead minnows (Pimephales promelas), and water fleas (Daphnia magna) (Karen et al., 1999). Furthermore, the higher concentration of calcium and magnesium in natural water may also be a factor to compete with test chemicals by reducing chemical binding at a target site, resulting in varied toxicity to D. carinata. This has been previously reported to reduce copper toxicity to Daphnia magna (De Schamphelaere and Janseen, 2002).

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Table 8.1 Physico-chemical properties of waters

Dissolved - -2 EC organic Cl SO4 Na Ca Mg K Fe Cu As Cr Cd Co Ni Water pH (μS) carbon (mg L-1) (µg L-1)

Cladoceran 6.7 66.4 0.38 14.0 0.98 7986 213 1229 618 3.97 1.00 0.40 nd 0.001 0.23 0.15 water

Morialta 7.0 651 4.02 106 15.3 69810 23650 23930 3782 36.9 2.07 1.05 1.10 0.06 0.14 2.57 Falls

Kangaroo 7.1 375 8.56 45.5 6.45 46320 23780 14510 6494 18.4 0.93 0.72 0.04 0.01 0.30 0.06 Creek

Note: Fluoride, nitrite, nitrate, phosphate, zinc, and lead were not detected in the water samples

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8.3.2 Acute toxicity of test chemicals The acute toxicity values of MAP and PSE to D. carinata is presented in Table 8.2.

The LC50 values of MAP to D. carinata in cladoceran water at 24 and 48 h were 36.3 and 21.8 -1 -1 mg L , respectively. Higher LC50 values were recorded for MAP (47.1 and 36.3 mg L ) at 24 h and 48 h in non-sterile KC natural water. The LC50 values of MAP to D. carinata in MF natural water (both non-sterile and sterile) were recorded higher than cladoceran water but less than KC natural water (42.5 and 39.8 mg L-1 at 24 h and 31.7 and 24.6 mg L-1 at 48 h, respectively). -1 The LC50 values of PSE to D.carinata in cladoceran water were 70.9 and 55.3 mg L , respectively, for 24 and 48 h whereas the corresponding values in non-sterile KC natural water were higher (104.3 mg L-1 at 24 h and 79.59 mg L-1 at 48 h, respectively). No appreciable -1 difference was recorded in the LC50 values of PSE between sterile water KC (78.6 mg L at 24 h and 65.8 mg L-1 at 48 h, respectively) and MF (77.8 mg L-1 at 24 h and 62.5 mg L-1 at 48 h, respectively). This reduction of toxicity in non-sterile natural waters could be due to the capacity for water suspended particles to physico-chemically interact with both organic and inorganic contaminants in water (Ran et al., 2000). The influence of water suspended particles on reducing chemical bioavailability and toxicity has been well documented for different compounds: hexachlorobenzene and lindane (Ekelund et al., 1987); pyrethroid insecticides (Yang et al., 2009); atrazine (Phyu et al., 2004); carbofuran to D. magna (Herbrandson et al., 2003) and chlordane to D. magna (Hall et al., 1986). MAP’s mode of action for causing mortality to D. carinata was not established. However, MAP toxicity in humans leads to neurodegenerative changes in the brain and in animals the degeneration of monoaminergic terminals and neuronal apoptosis (Krasnova and

Cadet, 2009). These stimulants demonstrated varied toxic responses (LD50) to animal models, for instance, intravenous (IV) administrations of 10 mg Kg-1 of amphetamine killed dogs within 3 h whereas oral administrations of amphetamine sulfate and methamphetamine were 20 - 27 and 9 -100 mg Kg-1, respectively (Volmer, 2005). In all test water samples, MAP and PSE were stable during the incubation period, which is consistent with the persistence of MAP in soil for a longer period of time (Pal et al., 2011). Microbial metabolites of MAP and PSE, therefore, can not be responsible for the observed mortality of D. carinata. Perhaps water quality might be the dominant factor controlling the toxicity of MAP and PSE. Hence, no specific conclusion can be made regarding the toxicity of MAP and PSE to D. carinata until more detailed research employing a range of waters with varying physio-chemical properties is undertaken.

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Table 8.2 Acute toxicity of methamphetamine and pseudoephedrine to Daphnia carinata

Methamphetamine (MAP) Pseudoephedrine (PSE) Water LC50 – 24 h Correlation LC50 – 48 h Correlation LC50 – 24 h Correlation LC50 – 48 h Correlation samples (mg L-1) coefficient (mg L-1) coefficient (mg L-1) coefficient (mg L-1) coefficient

Cladoceran 36.3 ± 1.39 0.98** 21.8 ± 1.02 0.95** 70.9 ± 3.19 0.92** 55.3 ± 1.91 0.96** water

Morialta Falls non-sterile 42.5 ± 1.51 0.98** 31.7 ± 1.31 0.98** 97.1 ± 4.00 0.97** 68.9 ± 2.85 0.93** water

Kangaroo Creek 47.1 ± 1.61 0.97** 36.3 ± 1.49 0.95** 104.3 ± 3.80 0.94** 79.6 ± 3.44 0.91** non-sterile water

Morialta Falls 39.8 ± 1.47 0.99** 24.6 ± 1.17 0.91** 77.8 ± 3.14 0.93** 62.5 ± 2.39 0.93** sterile water

Kangaroo Creek sterile 40.4 ± 1.52 0.98** 29.6 ± 1.27 0.97** 78.6 ± 3.24 0.94** 65.8 ± 2.76 0.98** water Figures in parenthesis represent standard error,*p ˂ 0.05; **p ˂ 0.00

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8.3.3 Genotoxicity of MAP and PSE to D. carinata The impact of genotoxic chemicals on the integrity of cellular DNA is one of the first events in organisms exposed to contaminants. The comet assay or single cell gel electrophoresis has broad applicability when applied to aquatic bio-indicator organisms, providing a sensitive, rapid and versatile system for the study of environmental genotoxicity (Frenzilli et al., 2009). Tail DNA percentage and olive tail movement were chosen for evaluating DNA damage of D. carinata and the results are shown in Figure 8.1. Compared to controls, test chemical MAP showed significant (p< 0.05) DNA damage and olive tail movement in 0.25, 0.5 and 1 mg L-1 concentrations whereas PSE caused damage at 0.5 and 1 µg L-1 concentrations. Figure 8.2a shows minimum DNA damage in D. carinata cells in controls, whereas MAP and PSE exposed daphnid cells exhibited DNA migration and damage as a result of strand breakage (Figure 8.2b and 8.2c). MAP concentrations between 5 and 100 µg L-1 and PSE concentrations ranging from 5 to 250 µg L-1 did not cause any significant DNA damage and olive tail movement to D. carinata, which suggest that the concentrations of MAP and PSE reported in the environment are not genotoxic. Although the lower experimental concentrations of these compounds are not genotoxic, as these compounds have potent pharmacological and biological activities and are present in water bodies together with the residues of many pharmaceuticals and other organic compounds the mixed effects could be unpredictable (Pal et al., 2013). In general, the most important damage at the molecular level is the induction of various types of lesions in DNA including strand breaks, which are efficiently measured by the comet assay (Jha, 2008). Such breaks give rise to chromosomal aberrations if they are un- or mis- repaired. Chromosomal aberrations can lead to cell death, which could lead to several pathophysiological conditions (Fadeel et al., 1999). Our results align with previous studies reporting DNA damages to vertebrates after acute and chronic exposure to MAP (Tokunaga et al., 2008). MAP is toxic to rat organs, including the brain, even at single lower dosage administration by elevating lipid oxidation products (Gluck et al., 2001; Açikgöz et al., 2000). Embryonic oxidative DNA damage was recorded in adult mice (Jeng et al., 2005). A few studies report that PSE induce apoptosis activities, renal tubular cell degeneration, and cytotoxicity to human neuroblastoma SH-SY5Y and rat myoblastoma H9c2 (2-1) cell lines (Nudmamud-Thanoi and Thanoi, 2012; Soni et al., 2004). The apoptosis and cytotoxicity activity are an indicator of cell death, necrosis and also the sign of DNA strand breakage or fragmentation (Brown and Attardi, 2005). Taken together, MAP and PSE can induce cell damage in daphnia’s cells, possibly as a result of oxidative DNA damage. However, while ephedrine influences various minor to fatal side-effects, short-term exposure do not induce

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genetic damages (Soni et al., 2004; Radaković et al., 2011). The present findings must be interpreted cautiously because the comet assay result does not necessarily predict the mutagenic effects of the MAP and PSE. Therefore, this result suggests that MAP and PSE can induce cell damages to daphnids cells. The measurement of genotoxic effects of MAP and PSE using genotoxicity biomarker tests such as comet assay in Daphnia could be a useful tool for monitoring aquatic toxicity due to illicit substances. Considering the potential of D. carinata as a bio-indicator species, the importance of the genotoxicity of MAP and PSE in ecotoxicity monitoring, the measurement of the DNA damage in this species after exposure to MAP and PSE could provide useful information for monitoring aquatic ecosystems where majority of illicit drugs finally ends up.

Figure 8. 1 DNA damage induced by MAP and PSE to D. carinata. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA & olive tail movement - ANOVA, p < 0.05)

12.0 1.60 * 1.40 10.0 % tail DNA h Olive tail movement 1.20 8.0 * 1.00

6.0 g 0.80 Olive tail movement tail Olive % tail DNA 0.60 4.0 * 0.40 f 2.0 0.20 cd cd e bc 0.0 a b 0.00 0 5 10 25 50 100 250 500 1000 MAP concentration (µg L-1)

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7.0 0.70 * 6.0 0.60 % tail DNA c Olive tail movement 5.0 0.50

4.0 0.40

3.0 0.30

% tail DNA * 2.0 0.20 b movement tail Olive 1.0 0.10

a a a a 0.0 a a a 0.00 0 5 10 25 50 100 250 500 1000 -1 PSE concentration (µg L )

Figure 8.2 DNA damage in D. carinata exposed to MAP and PSE. (a). Control with no or minimal DNA damage (b) DNA damage in MAP 1 mg L-1 exposed D. carinata and (c) DNA damage in PSE 1 mg L-1.

(a) (b) (c)

8.4 Conclusion To the best of our knowledge, this study is the first to reveal the acute and geno - toxicity of MAP and PSE to D. carinata. The overall acute toxicity of these test compounds is influenced by suspended particulate matter in water. The other aqueous constituent (DOC, pH, Cl-, Ca & Mg) was also a determining element, as it can potentially affect the toxicity of chemicals in waters (Phyu et al., 2005; Karen et al., 1999; de Schamphelaere & Janssen, 2002). The acute toxicity of MAP and PSE in water to D. carinata indicated the following trend: cladoceran medium water ˃ sterile natural water ˃ non-sterile natural water, and the genotoxicity results of MAP and PSE emphasize a potential risk to inhabitants of aquatic

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system. Fortunately, the environmental concentrations of MAP and PSE found in the environment are less than those found to be toxic to D. carinata in our experiments, therefore adverse environmental impacts should be unlikely. To adequately characterize the toxicity of MAP and PSE towards aquatic organisms and their consequential ecological implications, a number of investigations have to be carried out, specifically: (i) the delayed effects of chronic exposure to different aquatic organisms; (ii) behaviour of mixture of compounds and their possible impacts; (iii) in-depth research into the influence of water parameters on toxicity; (iv) mechanism of action of parent drug and metabolites to aquatic organisms; and (v) effects on other species of water fleas.

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8.5 References

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Frenzilli G, Nigro M, Lyons BP. The Comet assay for the evaluation of genotoxic impact in aquatic environments. Mutat Res 2009; 681:80-92. Gluck MR, Moy LY, Jayatilleke E, Hogan K A, Manzino L, Sonsalla PK. Parallel increases in lipid and protein oxidative markers in several mouse brain regions after methamphetamine treatment. J Neurochem 2001; 79:152-160. Hall W, Degenhardt L, Sindicich N. Illicit drug use and the burden of disease. In: Heggenhougen K, Quah S, editors. International encyclopaedia of public health. Elsevier 2008, 523-30. Hall WS, Dickson KL, Saleh FY, Rodgers JH. Effects of suspended solids on the bioavailability of chlordane to Daphnia magna. Arch Environ Contam Toxicol 1986; 15:529-534. Herbrandson C, Bradbury SP, Swackhamer DL. Influence of suspended solids on acute toxicity of carbofuran to Daphnia magna I. interactive effects. Aqu Toxicol 2003; 63:333- 342. Irvine RJ, Kostakis C, Felgate PD, Jaehne EJ, Chen C, White JM. Population drug use in Australia: A wastewater analysis. Forensic Sci Int 2011; 210:69-73. Jeng W, Wong AW, Ting-A-Kee R, Wells PG. Methamphetamine-enhanced embryonic oxidative DNA damage and neurodevelopmental deficits. Free Radi Biol Med 2005; 39: 317-326. Jha AN. Ecotoxicological applications and significance of the comet assay. Mutagenesis 2008; 23:207–221. Karen DJ, Ownby DR, Forsythe BL, Bills TP, La Point TW, Cobb GB, Klaine SJ. Influence of water quality on silver toxicity to rainbow trout (Oncorhynchus mykiss), fathead minnows (Pimephales promelas), and water fleas (Daphnia magna). Environ Toxicol Chem 1999; 18:63-70. Krasnova IN, Cadet JL. Methamphetamine toxicity and messengers of death. Brain Res Rev 2009; 60:379-407. Megharaj M, Singleton I, Mcclure N, Naidu R. Influence of petroleum hydrocarbon contamination on microalgae and microbial activities in a long-term contaminated soil. Arch Environ Contam Toxicol 2000; 38:439-445. Nikkila A, Paulasson M, Almgren K, Blankck H, Kukkonen JVK. Atrazine uptake, elimination, and bio-concentration by periphyton communities and Daphnia magna: effects of dissolved organic carbon. Environ Toxicol Chem 2001; 20:1003–1011. Nudmamud-Thanoi S, Thanoi S. Pseudoephedrine induces sperm abnormalities, lower sperm counts and increased apoptosis in rat testis. Cell Tiss Res 2012; 349:625-630. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Biotic and abiotic degradation of illicit drugs, their precursor, and by-products in soil. Chemosphere 2011; 85:1002-1009.

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Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Illicit drugs and the environment - A review. Sci Total Environ 2013; 463:1079-1092. Park SY, Choi J. Geno-and ecotoxicity evaluation of silver nanoparticles in freshwater crustacean Daphnia magna. Environ Engi Res 2010; 15(1):23-27. Phyu Y, Warne M, Lim R. Toxicity of atrazine and molinate to the cladoceran Daphnia carinata and the effect of river water and bottom sediment on their bioavailability. Arch Environ Contam Toxicol 2004; 46:308-315. Pomati F, Castiglioni S, Zuccato E, Fanelli R, Rossetti C, Calamari D. Effects of environmental contamination by therapeutic drugs on human embryonic cells. Environ Sci Tech 2006; 40:2442-7. Radaković M, Đelić N, Stanimirović Z, Plećaš-Solarović B, Spremo-Potparević B, Živković L, Bajić V. Evaluation of the effects of ephedrine on human lymphocytes in the comet assay. Acta veterinaria 2011; 61:363-371. Ran Y, Fu JM, Sheng GY, Beckett R, Hart B. Suspended particulate and colloidal matter in natural waters. J Environ Sci (China) 2000; 12:129-137. Singh NP, McCoy MT, Tice RR, Schneider EL. A simple technique for quantitation of low levels of DNA damage in individual cells. Expe Cell Res 1988; 175:184–191. Soni MG, Carabin IG, Griffiths JC, Burdock GA. Safety of ephedra: lessons learned. Toxicol Lett 2004; 150:97-110. Stojanovska N, Fu S, Tahtouh M, Kelly T, Beavis A, Kirkbride KP. A review of impurity profiling and synthetic route of manufacture of methylamphetamine, 3,4- methylenedioxymethylamphetamine, amphetamine, dimethylamphetamine and p- methoxyamphetamine. Forensic Sci Int 2013; 224(1-3):8-26. Tokunaga I, Ishigami A, Kubo SI, Gotohda T, Kitamura O. The peroxidative DNA damage and apoptosis in methamphetamine-treated rat brain. J Investi Med 2008; 55:241-5. UNODC (United Nations Office on Drugs and Crime). Amphetamines and Ecstasy: Global ATS Assessment. United Nations Publication; 2008.http://www.unodc.org/documents/scientific/ATS/Global-ATS-Assessment-2008- Web.pdf [Accessed on February 15, 2012]. UNODC (United Nations Office on Drugs and Crime).World Drug Report. United Nations Publication.2013. http://www.unodc.org/unodc/secured/wdr/wdr2013/World_Drug_Report_2013.pdf [Accessed on February 25, 2014]. van Nuijs ALN, Castiglioni S, Tarcomnicu I, Postigo C, de Alda ML, Neels H. Illicit drugs consumption estimations derived from wastewater analysis: a critical review. Sci Total Environ 2011; 409:3564-77.

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Vearrier D, Greenberg MI, Miller SN, Okaneku JT, Haggerty DA. Methamphetamine: history, pathophysiology, adverse health effects, current trends, and hazards associated with the clandestine manufacture of methamphetamine. Disease-a-month 2012; 58: 38-89. Volmer PA. Recreational drugs. In; Small Animal Toxicology. Peterson ME, Talcott PA (eds). Saunders, Philadelphia PA 2005; 2:273-311. Yang W, Spurlock F, Liu W, Gan J. Inhibition of aquatic toxicity of pyrethroid insecticides by suspended sediment. Environ Toxicol Chem 2009; 25:1913-1919. Zuccato E, Castiglioni S, Bagnati R, Chiabrando C, Grassi P, Fanelli R. Illicit drugs, a novel group of environmental contaminants. Wat Res 2008; 42:961–8

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CHAPTER 9

Acute and geno-toxicity of cocaine and 3, 4- methylenedioxymethamphetamine (MDMA) to Daphnia carinata

Abstract

Illicit drug abuse is considered as a global issue and these substances have been recognised as the latest emerging contaminants that can potentially cause environmental concerns particularly to aquatic biota. Although several studies have revealed the incidence of illicit drugs and their metabolites in different environmental systems worldwide at varying concentrations ranging from ng L-1 to µg L-1. The studies on the possible detrimental effects of these compounds to non-target organisms are inadequate. To expand the knowledge of illicit drugs toxicity to non-target aquatic organisms, the present study investigated the impact of acute and geno-toxicity of cocaine and 3, 4-methylenedioxymethamphetamine (MDMA) to fresh water cladoceran, Daphnia carinata. The study was conducted both in cladoceran and natural water collected from local creeks. The toxicity of both cocaine and MDMA followed the order: cladoceran culture medium ˃ sterile natural water ˃ non-sterile natural water. Cocaine and MDMA were found to be relatively less toxic in non-sterile natural water compared to sterile natural water samples, which could be due to the effect of varied water quality parameters of natural waters and also due to biodegradation. Liquid chromatography and mass spectroscopy revealed that cocaine was transformed to benzoylecgonine (BE) and ecgonine methyl ester (EME) in the all the test water samples. BE was found to be relatively stable and toxic to D. carinata as the parent compound. Cocaine exhibited significant genotoxicity to D. carinata at 5 µg L-1 - 1.0 mg L-1 and MDMA at 100 µg L-1 - 1.0 mg L-1 when compared to controls. Our results suggest that even sub lethal exposure of cocaine and MDMA contribute severe ill effects to D. carinata including genetic damage.

Keywords: Illicit drugs; cocaine; MDMA; Daphnia carinata; genotoxicity.

9.1 Introduction Currently illicit drugs are classified as emerging contaminants which attracted the attention of analytical and environmental chemists around the world due to their eco-toxic 143

nature (Parolini and Binelli, 2013). Worldwide, several illicit drugs monitoring studies revealed the existence of cocaine, amphetamine type-stimulants (ATS), 3, 4- methlenedioxymethamphetamine (MDMA), opioids (opium, heroin and prescription opioids), Δ9-tetrahydrocannabinol (THC), as well as their corresponding metabolites, in different environmental systems (wastewater, surface waters, sewage sludge and atmospheric air) even at ng L-1 concentration (Pal et al., 2013). The reported illicit drugs’ environmental concentration levels match those of common pharmaceuticals used for therapeutic purposes (Santos et al., 2010). A recent world drug report (UNODC, 2014) estimated that around 324 million people (3.5 -7.0% of the population) aged between 15-64) were found to have used illicit drugs at least once in 2012. Among the drugs, cannabis was the most widely used by 125 -227 million people followed by ATS, opioids and cocaine (UNODC, 2014). In Australia, use of methamphetamine (MAP), amphetamine and MDMA, remains the second most used illicit drug (40%) after cannabis (55%) whereas cocaine use was reported to be stable (11%) over the last decade. In Australia, a record number of ATS and cocaine seizures and drug- related arrests were occurred during 2012–13, which is the highest reported since the last decade (ACC, 2011-12). Cocaine is used by snorting (intranasally), injection (intravenously) or inhaling (smoked). In human cocaine causes sodium channel blocking, interferes with the uptake of epinephrine, serotonin and dopamine neurotransmitters, and vasoconstrictiveness (Binelli et al., 2012; EMCDDA, 2013; Gheorghe et al., 2008). Cocaine use in humans leads to various physiological effects such as stimulation of central nervous system, cardiac problems, addictiveness, pulmonary complication and serotonin level alteration (Binelli et al., 2012; Pal et al., 2013; Parolini and Binelli, 2013). Cocaine hydrolyses very rapidly in human system into benzoylecgonine (BE) and ecgonine methyl ester (EME) and some minor metabolites (Pal et al., 2013). Generally a negligible amount of cocaine (1-9%) is excreted as the parent compound through urine, but a major proportion of the substance is excreted in the form of its metabolites, BE (35-54%) and EME (32-49%) (Pal et al., 2013; Postigo et al., 2008) On the other hand, MDMA belongs to the amphetamine-type psychoactive compound which is structurally similar to amphetamines and mescaline (Lyles and Cadet, 2003). MDMA is commonly known as ecstasy and it is also referred under many different names such as love drug, dance drug, Adam, XTC, beans, hug, raves and E. Due to its hallucinogenic and stimulant effects, it is one the most popular recreational club drug prominent among young people aged between 15 and 34 (Hall and Henry, 2006; Pal et al., 2013). In recent years (2009 -2011) despite the decline in production and use of MDMA globally, unfortunately in Europe, North America and Oceania the trend is in positive magnitude (UNODC, 2014). It can be manufactured in clandestine laboratories using the following precursors, safrole, isosafrole,

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piperonal and 3, 4-methylenedioxyphenyl-2-propanone (PMK) and other locally accessible chemicals (EMCDDA, 2013). Most of the MDMA administered in humans is excreted in urine (26-65%) as a parent compound and some metabolites (Postigo et al., 2008). MDMA exposure in humans results in various outcomes such as brain microvasculature, white matter maturation, axonal damage (De Win et al., 2008). Furthermore, brain neurotoxicity damage leads to psychiatric, neuroendocrine and cognitive disorders (Gouzoulis-Mayfrank et al., 2006). Chronic use alters the cerebral blood flow or blood volume, amount of brain grey matter and working memory (Cowan, 2007). Despite several reports having been published on the presence of cocaine and MDMA in different environmental systems and their mammalian toxicity effects, reports on eco-toxicity of these compounds are lacking and there is no systematic information on potential detrimental effects on aquatic organisms (Binelli et al., 2012). To date only a very few reports are available on the eco-toxicity of amphetamine, cocaine, morphine and Δ-9-tetrahydrocannabinol (Pal et al., 2013; Parolini and Binelli, 2013) on aquatic organisms. The fresh water flea, D. carinata, found in rivers, creeks and fresh water lakes in Australia, and serve as sentinel organisms in the natural environment. There are a number of reasons for using D. carinata as a test organism, such as it is indigenous Australian fresh water species, and relatively larger than D. magna, has a high neonate production and is easy to handle (Phyu et al., 2004). This present study was designed to critically investigate the acute and geno-toxicity of cocaine and MDMA to D. carinata and the stability in both cladoceran and natural test water samples.

9.2 Materials and methods Refer chapter 8.2

9.3 Results and discussion

9.3.1 Physico-chemical properties of water The physico-chemical properties of test waters are presented in Table 8.1. Both natural water quality aspects varied noticeably from cladoceran water. MF and KC natural water had 4.0 and 8.6 mg L-1 of dissolved organic carbon (DOC) whereas cladoceran water contained less than 0.38 mg L-1. The electrical conductivity (EC) in MF and KC natural waters were 10- and 6-times greater than cladoceran water, respectively. Correspondingly, magnesium, sulfate, sodium and chloride content in MF and KC natural waters were 12-19, 7-16, 5-9, and

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3-8 times more than the cladoceran water, respectively. Fluoride, nitrite, nitrate, phosphate, zinc, and lead were not detected in natural waters.

9.3.2 Acute toxicity of test chemicals The acute toxicity values of cocaine and MDMA to D. carinata are presented in Table -1 9.1. The higher LC50 values 16.3 and 11.3 mg L to D. carinata was recorded in cocaine spiked cladoceran water at 24 and 48 h, respectively. The toxicity of cocaine was reduced in non- -1 sterile MF and KC natural water with the LC50 values of 39.1 and 29.4 mg L (24 h), and 38.8 -1 and 26.9 mg L (48 h). Both sterile natural water (MF and KC) recorded the same LC50 value -1 -1 (24.5 mg L ) at 24 h while at 48 h the LC50 values were 19.7 and 18.6 mg L , respectively.

The LC50 values of MDMA to D. carinata in cladoceran water at 24 and 48 h were 51.8 and

-1 Table 9.1 LC50 values (mg L ) of cocaine and MDMA to D. carinata tested in water samples

Water Cocaine MDMA samples 24 h 48 h 24 h 48 h

16.3 11.3 38.5 Cladoceran 51.8 (10.5 - 33.7) (5.9 - 18.7) (24.9 - 55.9) water (37.8 - 69.6)

Morialta Falls 39.1 29.43 85.7 79.7 non-sterile (25.1 - 51.4) (15.2 - 44.6) (67.5 - 117.0) (62.9 - 107.1) water

Kangaroo 38.8 26.97 79.7 68.8 Creek (24.1 - 50.5) (11.1 - 40.7) (62.9 - 107.1) (51.7 - 94.5) non-sterile

water

24.5 19.7 75.3 61.2 Morialta Falls (18.4- 36.8) (13.6 - 27.2) (59.1 - 99.4) (44.7 - 83.9) sterile water

Kangaroo 24.5 18.6 73.5 52.6 Creek (18.4 – 36.8) (12.5 - 26.1) (55.4 - 102.1) (37.1 - 73.9) sterile water

Values in parenthesis represents 95% confidence interval

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38.5 mg L-1, respectively, whereas the corresponding values for non-sterile MF natural water were lower (85.7 mg L-1 at 24 h and 79.7 mg L-1 at 48 h, respectively). Little difference was recorded in the LC50 values of MDMA between MF and KC sterile natural water samples at 24 -1 h, while at 48 h the LC50 values were 61.2 and 52.6 mg L , respectively. The toxicity reduction of cocaine and MDMA in non-sterile natural waters could be due to the sorption of suspended particles which may have reduced the chemical’s availability to D. carinata. Aqueous constituents such as DOC and suspended particulate matter in natural waters MF and KC might have reduced the bioavailability of cocaine and MDMA’s to D. carinata thereby reducing the toxicity. Earlier studies by Nikkila et al. (2001) and Phyu et al. (2004) demonstrated the marked effects of DOC and sediments on reducing the toxicity of organic pollutants to Daphnia magna. In addition, natural waters used in this study had slightly higher pH and high chloride along with DOC, which may decrease the chemical’s toxicity. Similar effects have been reported previously for silver toxicity to rainbow trout (Oncorhynchus mykiss), fathead minnows (Pimephales promelas), and water fleas (D. magna) (Karen et al.,1999). Furthermore, the higher concentration of calcium and magnesium in natural water may also be a factor that competes with test chemicals by reducing chemical binding at a target site, resulting in varied toxicity to D. carinata. This has been previously reported as having reduced the toxicity of copper on D. magna (De Schamphelaere and Janssen, 2002). Generally, the suspended particles have the capacity to interact physico-chemically with both organic and inorganic contaminants in water (Ran et al., 2000), which agrees with our results showing the least cocaine and MDMA toxicity to D. carinata in non-sterile natural water. In addition, based on the KOC value, cocaine (2.78) is expected to have low mobility which may influence the sorption of these compounds into suspended solids and sediment present in the water. Furthermore, the pKa value of cocaine (8.6) and MDMA (9.9) indicates that these compounds will exist partially in cation form in the environment. Generally, a cation can be adsorbed more strongly to organic carbon and clay components in soil, sediments and suspended particle compared to its neutral counterpart (NCBI). The influence of water suspended particles on reducing chemical bioavailability and toxicity has been well documented for different compounds: hexachlorobenzene and lindane (Ekelund et al., 1987); pyrethroid insecticides (Yang et al., 2009); atrazine (Phyu et al., 2004); carbofuran to D. magna (Herbrandson et al., 2003); and chlordane to D. magna (Hall et al., 1986). Reports by Phyu et al. (2004) and Petrie et al. (2014) on the suspended sediment’s sorption of chemicals resulting in reduced bioavailability and toxicity to aquatic organisms agree with our results. With particular reference to cocaine toxicity, its reduction in non-sterile natural water could be also due to the microbial mediated degradation, which is evident due

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the fact that 95.8% of the initially spiked cocaine disappeared in this water compared to less degradation occurring in the sterile natural water and cladoceran water.

In acute toxicity studies, the LC50 values can be categorised into three groups based on their effects: harmful between 10 and 100 mg L-1; toxic (1 -10 mg L-1); and very toxic (˂ 1 -1 mg L ) to aquatic biota (Cleuvers, 2003; Petrie et al., 2014). The LC50 value observed in this study for cocaine was (irrespective of the test water) between 11.3 and 39.1 mg L-1 and the corresponding value for MDMA was between 38.5 and 85.7 mg L-1. According to the above findings, cocaine and MDMA fall into the category of harmful to D. carinata. Illicit drugs’ LC50 value to aquatic organisms are limited which makes it difficult to compare our results with those of other studies. Most toxicity studies used a range of test species and varied their toxicological end points to measure the LC50 values (Petrie et al., 2014). However, some of the emerging contaminants such as non-steroidal anti-inflammatory drugs, lipid regulators and carbamazepine and trimethoprim LC50 values to aquatic organisms are comparable with our findings (Cleuvers, 2004; Petrie et al., 2014].

9.3.3 Stability of cocaine and MDMA In all test water samples, MDMA was relatively stable during 48 h incubation under the experimental conditions. An average of 97.5% of initial spike concentration (5 mg L-1) of MDMA was recovered at the end of the incubation period. However, the cocaine was found to be degraded at various levels in different test water samples. The recovery of cocaine in cladoceran water and sterile natural water was 64.1% and 49.7%, respectively, while the corresponding value in non-sterile natural water was very low (4.1 - 4.2%) at 48 h incubation (Table 9.2). A substantial reduction of cocaine concentration in the test water samples was accompanied by an increase in the cocaine metabolites, BE and EME. Gheorghe et al (2008) and Postigo et al (2008) have reported that cocaine was degraded very quickly in surface water (75% after 1 day) at 20 °C in comparison to HPLC grade water. This result suggests that degradation of cocaine in non-sterile natural water is mainly due to the microbial degradation process in addition to other processes. Additionally, chemical hydrolysis, enzymatic hydrolysis and photolysis processes have also been reported in accounts for cocaine transformation into its metabolites such BE and EME (Gheorghe et al., 2008; Postigo et al., 2008). Although cocaine disappeared very rapidly in non-sterile natural water, the main metabolite BE was observed to remain stable for a period of time (5 days), which agrees well with the Gheorghe et al. (2008) report on stability profiles of BE in surface and wastewater (pH 6) at 20 °C.

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Table 9.2 Stability of cocaine and MDMA in water samples

Initial nominal % chemical recovered after Water samples concentration 48 h incubation -1 (mg L ) Cocaine MDMA

Cladoceran water 5 64.1 ± 1.23 98.6 ± 0.77

Morialta Falls 4.2 ± 0.41 5 96.0 ± 0.65 non-sterile water

Kangaroo Creek 5 4.1 ± 0.37 97.0 ± 0.23 non-sterile water

Morialta Falls 5 49.7 ± 0.91 98.2 ± 0.13 sterile water

Kangaroo Creek 49.3 ± 1.21 5 97.8 ± 0.26 sterile water

9.3.4 Geno-toxicity of cocaine and MDMA to D. carinata The comet assay or single cell gel electrophoresis has broad applicability when applied to aquatic bio-indicator organisms, providing a sensitive, rapid and versatile system for the study of environmental genotoxicity (Frenzilli et al., 2009). Tail DNA percentage and olive tail moment were chosen for evaluating DNA damage of D. carinata and the results are shown in Figure 9.1. Cocaine showed significant (p< 0.05) DNA damage and olive tail moment in low, medium and high concentrations compared to controls whereas the corresponding effects were noticed in medium and higher concentrations (250 -1000 µg L-1) of MDMA exposed D. carinata. In general, the DNA damage and olive tail moment were found to be concentration- dependent with the highest DNA damage observed at a higher concentration, followed by medium concentration and low concentration. Even at a lower test concentration (5 µg L-1), cocaine proved to be more genotoxic than MDMA. The experimental lower concentration of cocaine (5 µg L-1) is comparable to some of the present environmental concentrations of cocaine and metabolite (BE) in wastewater influent (4 – 4700 ng L-1 and 9- 7500 ng L-1) (Pal et al., 2013). Figure 9.2a shows minimum DNA damage in D. carinata cells in the control, whereas cocaine- and MDMA-exposed daphnid cells exhibited DNA migration and damage as a result of strand breakage (Figure 9.2b and 9.2c). MDMA concentrations between 5 and 50 µg L-1 did not cause any significant DNA damage and olive tail movement to D. carinata, which suggests that the present environmental concentration of MDMA was not genotoxic.

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Though the lower concentrations of MDMA were not genotoxic, since these compounds have potent pharmacological effects on non-target organisms and their presence in water bodies together with the residues of many pharmaceuticals and other organic compounds, the mixed effects could be unpredictable (Pal et al., 2013). The results in the present study on the genotoxic effects of cocaine on daphnid cells is in agreement with those of previous studies on different organisms. Binelli et al. (2012) reported that an environmental concentration of cocaine induced significant primary DNA damage and increased the number of apoptosis cells in Dreissena polymorpha which clearly confirms the cyto-genotoxicity effects of cocaine in aquatic biota, mouse oocytes (Combelles et al., 2000), human foetuses (Meyer and Zhang, 2009). Cocaine metabolites have been observed to cause mitogenic effects in protozoan Tetrahymena pyriformis due to the mitotic stimulation process (Stefanidou et al., 2002). Generally, the mitogenic stimulation is harmful to cells and increases the DNA content. Alvarenga et al. (2011) recorded a clear primary DNA damage 1 h after exposure to cocaine (1.7– 7 mg kg-1 body weight in mice). The genotoxicity effects observed in our study could be due to the alteration of protein, which particularly was involved in segregation of the genome, mitotic spindle protein, cohesins (Binelli et al., 2012). Further, the observed genotoxic effects in our experiment could also be due to another possible mechanism of cocaine that increases oxidative stress. Generally, increases in oxidative stress can cause direct DNA damage, protein alteration and apoptosis (Binelli et al., 2012). Additionally, cocaine metabolites (BE and EME) have also induced similar effects such as primary DNA damage, oxidative stress and apoptotic process to D. polymorpha in environmental concentrations (Binelli et al., 2012). MDMA is mostly studied in the context of its effects on human and animal health in terms of specific neurotoxic and cardiac damage. The available literature on human and mammalian models exposed to MDMA cannot be compared for predicting the possible risks of MDMA to aquatic organisms, because the previously tested doses are much higher than our experimental concentrations. However, MDMA appeared to induce common pathway and cellular changes, and its pharmacological properties are similar to cocaine, considering that MDMA may also be dangerous to aquatic organisms in environmental concentrations. Additionally, the ability of MDMA to damage cells has been reported by several studies using rat liver and hippocampal cells (García-Cabrerizo and García-Fuster, 2014) and sperm cells (Barenys et al., 2009). Measuring the genotoxic effects of cocaine and MDMA employing a genotoxicity biomarker such as comet assay in Daphnia could be a useful tool for monitoring the aquatic toxicity of illicit substances. Considering the potential of D. carinata as a bio-indicator species, the importance of the genotoxicity of cocaine and MDMA in ecotoxicity monitoring, the

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measurement of the DNA damage in this species after exposure to cocaine and MDMA could provide useful information for monitoring aquatic ecosystems where the majority of illicit drugs end up.

Figure 9.1 DNA damage induced by cocaine and MDMA to D. carinata. Results are expressed as mean + SD.*p < 0.05 when compared to control and treatments. (% tail DNA & olive tail movement - ANOVA, p< 0.05)

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Figure 9.2 DNA damage in D. carinata exposed to cocaine and MDMA. (a). Control with no or minimal DNA damage (b).DNA damage in cocaine (1 mg L-1) exposed D. carinata (c) DNA damage in 1 mg L-1 of MDMA exposed D. carinata in water.

(a) (b) (c)

9.4 Conclusion The results obtained in this study reveal that cocaine and MDMA have clearly acute and genotoxic effects on D. carinata. Test waters varying in terms of quality parameters are believed to be the main factors that influence the acute toxicity of the test chemicals to D. carinata. The acute toxicity of cocaine and MDMA in water to D. carinata occurred in the following order: cladoceran medium water ˃ sterile natural water ˃ non-sterile natural water, and the genotoxicity results of cocaine and MDMA emphasise a probable risk to common non- target organisms. Comet assay in Daphnia could serve as a useful tool for monitoring the genotoxic effects of illicit drugs in water bodies. However, a comprehensive investigation should be undertaken to generate more substantive conclusions on cocaine and MDMA’s ecological implications. In-depth toxicity mechanism of these compounds and their metabolites should be further explored, because they could differ from those already established for human and biological models. Finally, understanding the effects of a mixture of compounds and their acute and chronic effects on different aquatic organisms will be essential for proper and thorough evaluation of the ecological effects of these compounds.

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9.5 References

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Frenzilli G, Nigro M, Lyons BP. The Comet assay for the evaluation of genotoxic impact in aquatic environments. Mutat Res 2009; 681:80-92 García-Cabrerizo R, García-Fuster MJ. Acute and chronic effects of MDMA on hippocampal cell fate markers in adolescent and adult rats. Eur Neuropsychopharmacol 2014; 24(2): 682-683. Gheorghe A, van Nuijs A, Pecceu B, Bervoets L, Jorens PG, Blust R. Analysis of cocaine and its principal metabolites in waste and surface water using solid-phase extraction and liquid chromatography-ion trap tandem mass spectrometry. Anal Bioanal Chem 2008; 391:1309-19. Gouzoulis‐Mayfrank E, Daumann J. Neurotoxicity of methylenedioxyamphetamine (MDMA; ecstasy) in humans: how strong is the evidence for persistent brain damage? Addiction 2006; 101:348-361. Hall A, Henry J. Acute toxic effects of ‘Ecstasy’ (MDMA) and related compounds: overview of pathophysiology and clinical management. Br J Anaesth 2006; 96:678-68. Hall WS, Dickson KL, Saleh FY, Rodgers JH. Effects of suspended solids on the bioavailability of chlordane to Daphnia magna. Arch Environ Contam Toxicol 1986; 15: 529-534. Herbrandson C, Bradbury SP, Swackhame, DL. Influence of suspended solids on acute toxicity of carbofuran to Daphnia magna I. Interactive effects. Aqua Toxicol 2003; 63: 333-342. Jones JI, Murphy JF, Collins AL, Sear DA, Naden PS, Armitage PD. The impact of fine sediment on macro-invertebrates. River Res Appl 2012; 28:1055-1071. Karen DJ, Ownby DR, Forsythe BL, Bills TP, La Point TW, Cobb GB, Klaine SJ. Influence of water quality on silver toxicity to rainbow trout (Oncorhynchus mykiss), fathead minnows (Pimephales promelas), and water fleas (Daphnia magna). Environ Toxicol Chem 1999; 18:63-70. Lyles J, Cadet JL. Methylenedioxymethamphetamine (MDMA, Ecstasy) neurotoxicity: cellular and molecular mechanisms. Brain Res Rev 2003; 42:155-168. Meyer KD, Zhang L. Short- and long-term adverse effects of cocaine abuse during pregnancy on the heart development. Ther Adv Cardiovasc Dis 2009; 3: 7-16. NCBI (National centre for biotechnology information) Cocaine compound Summary for: CID 446220. Environmental fate and exposure potential. http://pubchem.ncbi.nlm.nih.gov/rest/chemical/cocaine#x351 (Accessed on 15th October 2014).

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Nikkila A, Paulasson M, Almgren K, Blankck H, Kukkonen JVK. Atrazine uptake, elimination, and bio-concentration by periphyton communities and Daphnia magna: effects of dissolved organic carbon. Environ Toxicol Chem 2001; 20:1003-1011. Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Illicit drugs and the environment—A review. Sci Total Environ 2013; 463-464:1079-1092. Parolini M, Binelli A. Adverse effects induced by ecgonine methyl ester to the zebra mussel: A comparison with the benzoylecgonine. Environ Poll 2013; 182:371-378. Petrie B, Barden R, Kasprzyk-Hordern B. A review on emerging contaminants in wastewaters and the environment: Current knowledge, understudied areas and recommendations for future monitoring. Water Res http://dx.doi.org/10.1016/j.watres.2014.08.053 Phyu Y, Warne M, Lim R. Toxicity of atrazine and molinate to the cladoceran Daphnia carinata and the effect of river water and bottom sediment on their bioavailability. Arch Environ Contam Toxicol 2004; 46:308-315. Postigo C, de Alda MJL, Barceló D. Analysis of drugs of abuse and their human metabolites in water by LC-MS2: a non-intrusive toll for drug abuse estimation at the community level. Trends Anal Chem 2008; 27:1053-69. Ran Y, Fu J-M, Sheng G-Y, Beckett R, Hart B. 2000. Suspended particulate and colloidal matter in natural waters. J Environ Sci (China) 2000; 12:129-137. Santos LHMLM, Araújo AN, Fachini A, Pena A, Delerue-Matos C, Montenegro MCBSM. Ecotoxicological aspects related to the presence of pharmaceuticals in the aquatic environment. J Haz Mat 2010; 175:45-95. Stefanidou M, Chatziioannou A, Livaditou A, Rellaki A, Alevisopoulos G, Spiliopoulou H, Koutselinis A. DNA toxicity of cocaine hydrochloride and cocaine freebase by means of DNA image analysis on Tetrahymena pyriformis. Biol Pharm Bull 2002; 25:332-334. UNODC (United Nations Office on Drugs and Crime). World Drug Report. United Nations Publication.2014. http://www.unodc.org/unodc/secured/wdr/wdr2014/World_Drug_Report_2014.pdf [Accessed on September 25, 2014]. Yang W, Spurlock F, Liu W, Gan J. Inhibition of aquatic toxicity of pyrethroid insecticides by suspended sediment. Environ Toxicol Chem 2009; 25:1913-1919

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CHAPTER 10

Phytotoxicity of illicit drugs to Lemna minor L.

Abstract

This study assessed the toxic effects of commonly abused illicit drugs and a precursor using an aquatic plant (Lemna minor L.) as a test organism. Growth parameters were assessed according to standardised protocol ISO, 20079 and OECD 221 guidelines, 2006. The suitability of tissue accumulation and certain biochemical changes such as chlorophyll and proline content serving as biomarkers for illicit drugs aquatic toxicity was evaluated. Results revealed that growth parameters (frond numbers, fresh weight and relative growth rate) and biochemical (chlorophyll and proline) content were affected with the increase in the concentration of test chemicals. The toxicity of these compounds was as follows: cocaine ˃ methamphetamine (MAP) ˃ 3, 4-methylenedioxymethamphetamine (MDMA) ˃ pseudoephedrine (PSE). We found that of the growth parameters, fresh weight was the most appropriate indicator for the effects of illicit drugs. Obtained results demonstrate the usefulness of L. minor L. firstly asan illicit drugs’ aquatic toxicity indicator and secondly, the significance of selected biological parameters in the reliable assessment of phytotoxic potential of complex aquatic systems.

Keywords: Methamphetamine; Pseudoephedrine; MDMA; Cocaine; Lemna minor L.

10.1 Introduction Drug abuse is a global problem and it causes serious social, economic and health damage to the community (Pal et al., 2013). A recent world drug report shows that the overall global production and consumption of most illicit drugs has risen, and may continue to increase in the future (UNODC, 2014). The commonly abused illicit drugs are cannabis, amphetamines, opiates, and lysergic acid diethylamide (LSD) (Pal et al., 2013). In Australia the use of amphetamine-type stimulants (ATSs) is on the rise and this includes the amphetamines group (e.g., amphetamine, methamphetamine) and ecstasy group (e.g., 3, 4- methylenedioxymethamphetamine (MDMA) and analogous compounds) (UNODC, 2008). A record number of ATS, cocaine and pseudoephedrine (PSE) seizures and arrests were reported in 2012–13 in Australia, which is the highest reported in the last decade (ACC, 2012- 13).

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The amphetamines group accounts for more than three quarters of ATSs and currently they demand the most attention of all the synthetic illicit drugs (UNODC, 2008). In particular methamphetamine (MAP) is the most commonly produced ATS, because it can be easily manufactured in clandestine laboratories using locally available chemicals and precursors (Pal et al., 2013). In recent years, PSE has been used as one of the preferred precursors for illegal manufacturing of MAP which can be found in cold-flu and nasal decongestant over-the-counter medicines (Moser and Rayburn, 2007). Similar to MAP, MDMA can also be manufactured in clandestine laboratories using various precursors such as safrole, isosafrole, piperonal and 3, 4-methylenedioxyphenyl-2-propanone (PMK) and other locally available chemicals (EMCDDA, 2013). Although the use and production of MDMA has decreased in recent years (2009-2011), it has grown specifically in Europe, North America and the Oceania region (UNODC, 2014). Cocaine remains very popular and after cannabis, it is the second most used illicit drug in European countries (Binelli et al., 2012). Drugs that are consumed by people are mostly excreted via urine as a parent compound and their metabolites, and these substances end up in wastewater treatment plants where they are often incompletely removed and/or escape from treatment processes (Pal et al., 2013). Several researchers worldwide have demonstrated the presence of these substances not only in wastewaters, but also in the receiving surface waters, which may have implications for humans and aquatic biota (Pal et al., 2013; Pomati et al., 2006). Worldwide, the cocaine and ATSs concentration in surface waters ranges from ˂LOQ (limit of quantification) to hundreds of ng-1 (Pal et al., 2013). Most of these substances possess a potent pharmacological effect, and their occurrence in surface waters could produce an unpredictable interaction with other related compounds (Zuccato et al., 2008). The environmental impact of these potentially toxic chemicals is increasingly being recognized as a critical issue of concern. However, most previous studies on illicit drugs focussed on sewage epidemiology, and only very limited data are available on the acute and chronic toxicity of these compounds on the aquatic biota (Binelli et al., 2012). Ecosystem functioning directly relies on a range of organisms and plants which are very vital because they serve as primary producers, providing food sources for varieties of aquatic life, nutrient recycling and soil stabilization, and providing habitat for aquatic biota (Gubbins et al., 2011). If any adverse effects of illicit drugs upon plant growth can cause significant changes in the ecosystem functioning, the result may well be permanent damage. However, scientific information on the effects of illicit drugs on plants is non-existent. In this study, Lemna minor L. was used as a test plant. It is commonly found in fresh water and brackish ecosystems in temperate climates. Additionally this plant is an important food source and it provides habitat for various aquatic life forms. It has been used very frequently in

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ecotoxicological research as a representative of higher aquatic plants. Furthermore, L. minor L. is easy to culture and handle, have a high multiplication rate and reproduces vegetatively, and is highly sensitive to different contaminants. This makes L. minor L. a suitable aquatic test organism (Drost et al., 2007; Naumann et al., 2007). We therefore undertook a study with Lemna to: (a) determine potential harmful effects of illicit drugs on L. minor L. growth parameters; (b) investigate biochemical changes such as chlorophyll and proline content; and (c) evaluate stability and bioaccumulation of illicit drugs.

10.2 Materials and methods 10.2.1 Chemicals Refer chapter 3.2.1 10.2.2 Cultivation and growth of L. minor L. L. minor L. was obtained from the culture maintained at the Centre for Environmental Risk Assessment and Remediation, University of South Australia, Mawson Lakes campus, South Australia. The plants were grown in open glass containers in sterilized Stenberg -1 -1 -1 medium (3.46 mmol L KNO3, 1.25 mmol L Ca(NO3)2.4H2O, 0.072 mmol L K2HPO4, 0.66 -1 -1 -1 -1 mmol L KH2PO4, 0.41 mmol L MgSO4.7H2O, 1.94 µmol L H3BO3, 0.63 µmol L -1 -1 -1 ZnSO4.7H2O, 0.18 µmol L Na2MoO4.2H2O, 0.91 µmol L MnCl2.4H2O, 2.81 µmol L -1 FeCl3.6H2O, 4.03 µmol L EDTA Disodium-dihydrate; pH 5.5) in a controlled condition at 24 ± 2 °C (ISO/DIS 20079, 2004; OECD, 2006). L. minor L. fronds were surface sterilized using 0.5% (v/v) sodium hypochlorite solution for 1 minute and then rinsed with sterile deionized water in order to reduce the pathogen load. Surface sterilized fronds were grown for 14 days in Steinberg medium for acclimatization. Three to four frond colonies (5 colonies per treatment) (Naumann et al., 2007) were randomly selected and transferred to the test medium containing illicit drugs at concentrations ranging from 0 to 400 mg L-1. The test containers were incubated at 24 ± 2 °C in continuous cool white fluorescent lighting (6500 – 10000 lux) for an exposure duration of 7 days (7d) (standard OECD requirements). All exposures were executed in triplicate.

10.2.3 Growth parameters According to OECD guidelines (2006) for whole organism measurement of plant health, a) frond number was recorded before exposure and at the end of the exposure; b) fresh weight: each colony was collected and rinsed with distilled water and excess moisture was removed using tissue paper and weight was recorded; c) Relative Growth Rate (RGR) of the fronds was calculated according the standard OECD formula, (2006) (Eq. (1); d) inhibition

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of growth on the basis of frond number and fresh weight was calculated according to OECD guidelines (Eq. (2)).

( ) ( ) = ……………………………………………………...………….. (1) ln 𝑁𝑁𝑁𝑁 −ln 𝑁𝑁𝑁𝑁 where:𝜇𝜇𝑖𝑖−𝑗𝑗 𝑡𝑡𝑗𝑗− 𝑡𝑡𝑖𝑖 : average specific growth rate from time i to j

𝜇𝜇𝑖𝑖− 𝑗𝑗: measurement variable in the test or control vessel at time i 𝑁𝑁𝑁𝑁 : measurement variable in the test or control vessel at time j 𝑁𝑁𝑁𝑁 : time period from i to j 𝑡𝑡

% = × 100………………………………………………………………… (2) 𝜇𝜇𝑐𝑐− 𝜇𝜇𝑡𝑡 where:𝐼𝐼𝑟𝑟 𝜇𝜇𝑐𝑐 % : percent inhibition in average specific growth rate

𝐼𝐼 𝑟𝑟: mean value for µ in the control 𝜇𝜇𝑐𝑐 : mean value for µ in the treatment group 𝜇𝜇𝑡𝑡 10.2.4 Chlorophyll and free proline estimation L. minor L. exposed to illicit drugs for 7 days were removed from the vessels and rinsed with deionized water and dried with soft paper, and weighed in pre-weighed microcentrifuge tubes on ice. L. minor L. fresh tissue (125 mg) was homogenized with 80% (w/v) ice cold acetone, centrifuged at 5000 g for 10 min and the absorbance of the clear extract was measured using a spectrophotometer at 663, 646 and 470 nm (Synergy™ HT, Bio-Tek equipped with KC4 software). The chlorophyll content was estimated (mg g-1 fresh weight) from the equation (Lichtenthaler, 1987). Free proline content in the fronds was measured as described by Bates et al. (1973). Fresh tissue (125 mg) was manually homogenized in 3% (w/v) sulphosalycylic acid and centrifuged at 1000 g for 3 min. Ninhydrin reagent was added to the supernatant and heated at 100 °C for 1 h in water bath and cooled in ice. The chromophore obtained was extracted from liquid phase with toluene and the organic layer’s absorbance was read at 520 nm in the spectrophotometer (Synergy™ HT, Bio-Tek equipped with KC4 software). Proline concentration was determined from the calibration curve using L-proline as standard and expressed as nmol g-1 fresh weight.

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10.2.5 Bioaccumulation of illicit drugs In order to estimate the accumulation of illicit drugs, 10 mg L-1 exposed L. minor L. fronds (250 mg) were transferred into plastic tubes. Before being transferred into the plastic tubes fronds were surface washed and excess water was removed using filter paper. Fronds were manually homogenized in 100 mM PBS buffer and extracted with 10 ml of chloroform: acetonitrile: methanol: acetic acid (80:10:9:1) in a glass vial. This extraction was carried out in the following sequence: the sample was ground using a glass homogeniser (3- 5 min); vortexed (30 sec); electric shaker (30 min); and sonicated for 5 min @ 30 °C. The vials were centrifuged and the aliquots were filtered through 0.22 µm Teflon filters. The supernatants were dried and reconstituted in 1 ml of HPLC grade of methanol for HPLC-MS analysis.

10.2.6 Chemical analysis Refer chapter 4.2.4

10.3 Results and discussion

10.3.1 Effect of illicit drugs on growth parameters (frond number, RGR and fresh weight) Our results provide evidence that the growth of L. minor L. was significantly affected by illicit drugs and a precursor. Figures 10.1 and 10.3 present the total frond number and relative growth rate of L. minor L. over 7d in the presence of various test chemical concentrations. Cocaine and MAP indicated a significant (p < 0.05) effect on L. minor L. total frond number in low, medium and high concentrations compared to controls whereas the corresponding effect was noticed in medium and higher concentrations of MDMA and PSE exposed L. minor L. In general, the growth of L. minor L. proved to be concentration-dependent in all test chemicals. Negative growth was observed in L. minor L. at 300 and 400 mg L-1 of cocaine and MAP 7d exposure. Approximately half of the fronds were reduced at 10 mg L-1 of cocaine and at 25 mg L-1 of MAP and MDMA exposure while the corresponding value was observed only in ˃300 mg L-1 of PSE exposure. Relative growth rates (RGR) were calculated using the frond numbers which also indicated an effect on growth in 7d exposure to test chemicals. Cocaine highlighted a significant reduction (p < 0.001) on RGR compared to control at 10 mg L-1 (Figure 10.3a) which was evident after 3 days exposure, whereas MAP and MDMA produced such effects at higher concentrations (300 and 400 mg L-1). PSE did not cause any significant reduction in RGR at all concentrations in 7d exposure. A significant negative growth rate was observed at ˃ 300

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mg L-1 (cocaine and MAP) and ˃ 400 mg L-1 (MDMA) 7d treatments to Lemna. There were notable visual toxicity symptoms such as chlorosis, necrosis, and breaking of fronds and roots were observed at 100 mg L-1 and 200 mg L-1 of MAP and MDMA treatments, but not to the same extent as that observed at higher concentrations. All the above mentioned symptoms were noticed in low, medium and high concentrations of cocaine treatments. At medium concentration of MAP and MDMA (25 mg L-1 and 50 mg L-1, respectively) the number of fronds declined as the exposure days progressed (3d - 7d), suggesting that increased toxicity occurs over time. Overall, the cocaine exhibited an immediate response to RGR after 3 days of treatment compared to other test chemicals.

Figure 10.1 L. minor L. growth (total frond number) after exposure to cocaine, MAP, MDMA

and PSE for 7d. Bars denote standard deviation (n = 3). Different letters indicate a significant

difference (1-way ANOVA, p < 0.05).

After 7d exposure to cocaine, MAP and MDMA the fresh weight of L. minor L. was reduced significantly (1-way ANOVA, P ˂ 0.05) at 1 mg L-1 compared to the control, while PSE up to 25 mg L-1 did not cause any significant weight loss (Figure 10.2). Cocaine and MAP at 10 mg L-1 after 7d exposure was reduced to approximately half of the fresh weight of Lemna followed by a complete weight loss being recorded at ˃300 mg L-1. MDMA and PSE did not

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exhibit a complete weight loss in any of the test concentrations, but at 25 mg L-1 (MDMA) and 300 mg L-1 (PSE) weight loss of nearly half was observed.

Figure 10.2 L. minor L. growth (fresh weight) after exposure to cocaine, MAP, MDMA and

PSE for 7d. Bars denote standard deviation (n = 3). Different letters indicate a significant

difference (1-way ANOVA, p < 0.05).

Illicit drugs such as cocaine, MAP and MDMA are widely abused and their residues can be easily translocated into aquatic environments primarily through treated wastewater and also by manufacturers disposing of waste illegally into water bodies (Pal et al., 2013). The growth inhibition effects of these contaminants on non-target organisms requires certain criteria to differentiate their lethal and sub-lethal effects on plants. Using the following growth parameters such as frond number, RGR and fresh weight, may be of great relevance for evaluating the toxicity of illicit drugs on L. minor L. In general, the impacts of the tested chemicals were identified by the overall growth inhibition and biochemical alterations to L. minor L. From our study, it appeared that cocaine and MAP were more highly toxic than MDMA and PSE to L. minor L., and our results also revealed that all the growth parameters tested in 162

this experiment were equally sensitive to illicit drugs. Thus, L. minor L. can be used as a biomonitoring plant in aquatic systems to assess the illicit drugs’ toxicity.

Figure 10.3 L. minor L. growth (relative growth rate) after exposure to A) Cocaine B) MAP C)

MDMA and D) PSE for 7d. Bars denote standard deviation (n = 3). Different letters indicate a significant difference (1-way ANOVA p < 0.001).

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Several studies have used the L. minor L. due to above parameters to assess the toxicity of different chemicals in aquatic systems: heavy metals (Appenroth et al., 2010; Drost et al., 2007; Naumann et al., 2007); silver nanoparticles (Gubbins et al., 2011); polycyclic aromatic hydrocarbon (Zezulka et al., 2013); industrial effluent (Radić et al., 2010); and herbicides (Mitsou et al., 2006). They have suggested that these parameters are more or less equally sensitive to test compounds in comparison to other parameters tested in their studies. However, Radić et al. (2010) suggested that the frond number is considered to be the least reliable when comparing other growth parameters observed in L. minor L. This is possibly due to the fact that frond count is inappropriate to fresh weight. Additionally, plants may produce small buds when they are chemically stressed; these buds are often counted as individual fronds (Naumann et al., 2007; Radić et al., 2010). In fact, direct weighing of a sample may provide more reliable data in comparison to other parameters such as dry weight measurement and counting frond numbers. In our study, frond number was slightly less sensitive than fresh weight which agrees with the findings by Naumann et al. (2007) concerning heavy metals’ inhibition on L. minor L.

10.3.2 Effect of illicit drugs on inhibition of growth Statistical analysis to see the differences are not possible because the inhibition of growth is calculated using mean values of frond numbers and fresh weight. It is clear that inhibition of growth occurred at all concentrations of cocaine, MAP and MDMA while PSE at 165

lower concentrations (1 mg L-1 and 5 mg L-1) revealed very negligible effects (Figure 10.4d). Of these test chemicals, cocaine demonstrated the highest toxicity with 100% inhibition of growth found at ˃ 300 mg L-1 after 7d based on both frond number and fresh weight. The corresponding effect in MAP was observed at ˃ 200 mg L-1 (fresh weight) and at ˃ 400mg L-1 (frond number). Overall, the inhibition of growth in both (frond number and fresh weight) steadily increased with concentration with a few exceptions (Figure 10.4a-d). In our results, the growth parameters (frond number and fresh weight) exhibited different responses

regarding the inhibition of growth in all the test chemicals EC50 values were calculated for both frond numbers and fresh weight and found to be less for cocaine and more for PSE, with fresh weight being more greatly affected than frond numbers.

Figure 10.4 L. minor L. growth (inhibition of growth based upon fresh weight and frond number) after exposure to A) Cocaine, B) MAP, C) MDMA, D) PSE for 7d

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It appears that fresh weight was sensitive to cocaine, MAP and MDMA while frond number was sensitive to PSE. The reasons for these different responses are not known. In general, the EC50 of values in 7d exposure was recorded in the following descending order: PSE ˃ MDMA ˃ MAP ˃ cocaine (Table 10.1). The higher inhibition of cocaine to L. minor L. is believed to be linked to its quick transformation into benzoylecogonine (BE) and ecgonine methyl ester (EME) which have been observed to be toxic to the aquatic community (Rosi- Marshall et al., 2015). In our chemical analysis the results revealed the above cocaine metabolites were present in L. minor L. tissues as well as in the medium. Furthermore, the lower inhibition of PSE to L. minor L. may be related to its virtual disappearance from both the tissue and medium.

-1 Table 10.1 EC50 values (mg L ) of illicit drugs to L. minor L. based on growth measured as fresh weight and frond number (7d)

Drugs Fresh weight Frond number

Cocaine 8.92 10.03

MAP 8.75 37.52

MDMA 20.11 38.54 PSE 259.23 282.33

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10.3.3 Illicit drugs tissue accumulation Results for tissue accumulation of cocaine, MAP, MDMA and PSE at 10 mg L-1 concentration to L. minor L. over 7d exposure are presented in Table 10.2. Of all the test chemicals, MDMA recorded the highest accumulation (0.89 µg L-1) followed by MAP (0.71 µg L-1) and cocaine (0.54 µg L-1) while PSE accumulated the very least amount (0.04 µg L-1). In this test, measured concentrations were employed for making calculations. There were not many changes (less than 10%) observed between the test chemical’s nominal and measured concentrations in the medium. MAP and MDMA were stable during the test periods between 0 and 7d (Table 10.2), but cocaine and PSE were degraded, approximately 83% and 96%, respectively. The presence of L. minor L. in the test medium did not greatly affect the test chemicals’ concentration during the 7d period. Although these compounds were accumulated in very small amounts in L. minor L., tissues could be enough to cause a potential risk to the aquatic community. Since L. minor L. is an important food source for various aquatic life forms (Drost et al., 2007) which possibly make them to enter into the food chain. Additionally, cocaine, MAP and MDMA are biologically active, and consequently low level chronic exposure of these compounds is dangerous to aquatic biota (Pal et al., 2013). For example, cocaine and its metabolites BE and EME pose dangers to Dreissena polymorpha’s (Zebra mussel) physiology at environmental concentrations (Binelli et al., 2012; Parolini and Binelli, 2013; Parolini et al., 2013). Furthermore, accumulation of ATS such as MDMA and MAP may inhibit the growth of L. minor L. because the organic amines have been known to inhibit the biochemical and physiological process in algae (Crist et al., 1992; Rosi-Marshall et al., 2015). Table 10.2 Illicit drug concentration in medium and in L. minor L. tissues, exposed in 10 mg L-1 for 7d. Measured concentration in medium (mg L-1) Tissue Drugs accumulation 0 day 7 day (µg g-1) With Lemna No Lemna

Cocaine 9.97 ± 0.03 1.67 ± 0.08 1.89 ± 0.02 0.54 ± 0.02

MAP 10.04 ± 0.09 9.19 ± 0.03 9.85 ± 0.04 0.71 ± 0.03

MDMA 10.07 ± 0.02 8.98 ± 0.06 9.20 ± 0.07 0.89 ± 0.02

PSE 9.95 ± 0.02 0.39 ± 0.04 0.54 ± 0.08 0.04 ± 0.01 Data represent mean ± standard deviation of 3 replicates

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10.3.4 Chlorophyll and proline content The total chlorophyll content after 7d exposure to cocaine, MAP, MDMA and PSE is shown in Table 10.3. Illicit drugs had significantly decreased the chlorophyll content compared with the non-drugs control. This reduction was more pronounced as the drugs concentration increased. Cocaine and MAP significantly reduced the chlorophyll content at 1 mg L-1 and 5 mg L-1 exposure while MDMA and PSE exhibited similar significant effects at 100 mg L-1 and 300 mg L-1, respectively. MDMA and PSE concentrations between 1 and 50 mg -1 and 1 and 200 mg L-1 did not show any significant effect on L. minor L. chlorophyll contents. The results confirmed that the following order of test chemicals reduced the chlorophyll content: cocaine ˃ MAP ˃ MDMA ˃ PSE. The reduction of chlorophyll pigments is a key indicator of various noxious substances wielding toxic effects (Zezulka et al., 2013). In this study, we noticed chlorotic symptoms emerging in illicit drugs treatments after 4d and the severity of these symptoms increased over time and concentration. This may be related to decreased chlorophyll content or immobilization of photosynthetic nutrients caused by illicit drugs cation nutrient complex. In fact, plants require certain nutrients for chlorophyll formation and photosynthesis (Zobiole et al., 2011). Generally, plant biomass (fresh weight) production is determined mainly by the function of photosynthetic apparatus (Delgado et al., 1992). The significant changes in growth parameters (Figures 10.1- 10.4) of L. minor L. treated by illicit drugs correlate with the significant reduction in total chlorophyll content (Table 10.3). Furthermore, the chlorophyll content reduction was observed to be more pronounced in young fronds than in fully adult inoculated mother fronds. Hence, the synthesis of chlorophyll might constitute a possible target by illicit drugs and caused the documented outcomes. It can be speculated that the strong ability of illicit drugs to induce the formation of reactive oxygen species (Eskandari et al., 2013; Peraile et al., 2013) possibly indirectly inhibited the photosynthesis of L. minor L., and consequently oxidized chlorophyll. The reactive oxygen defence system is usually less developed in younger fronds than adult ones (Appenroth et al., 2010). In addition, it is well known that organic amines, aliphatic amines and amino acids which include certain illicit drugs can inhibit photosynthesis, nitrogen fixation and plants’ and algae’s physiological process (Rosi-Marshall et al., 2015). To the best of our knowledge, to date no reports have been published on illicit drugs, specifically comparing and discussing which mechanisms may be behind the chlorophyll reduction in L. minor L. Compared to the control, proline content was significantly increased in the treatments (Table 10.4). Cocaine, MAP and MDMA exhibited a significant effect at 25 mg L-1 while PSE did so at 100 mg L-1. A more significant effect was observed when test chemicals were in concentrations above 300 mg L-1. This result suggests that illicit drugs exposure to L. minor L.

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may have induced the stress and in turn may have increased the proline content in L. minor L. In general, proline accumulates in plant systems in response to abiotic stresses (Filippou et al., 2014) which regulate various enzymatic and physiological functions to protect plants from such stresses (Ashraf and Foolad, 2007; Spoljarevic et al., 2011). Similar results have been reported previously on various plants being affected by different abiotic stimuli: cadmium on maize seedlings and Atriplex halimus (Nedjimi and Daoud, 2009; Voetberg and Sharp, 1991); heavy metals on various field crops (Alia and Saradhi, 1991); chromium on Ocimum tenuiflorum (Rai et al., 2004); and drought stress on Ailanthus altissima (Filippou et al., 2014). While the stress response mechanisms vary from species to species and at different developmental stages, the basic cellular responses to abiotic stresses are unchanged among most plant species (Ashraf and Foolad, 2007).

Table 10.3 Chlorophyll contents after 7d exposure of L. minor L. to illicit drugs

Concentration Total chlorophyll content -1 (mg L-1) (mg g fresh weight) Cocaine MAP MDMA PSE

0 0.80 ± 0.02a 0.80 ± 0.02a 0.80 ± 0.02a 0.80 ± 0.02a

1 0.55 ± 0.01b 0.78 ± 0.06a 0.80 ± 0.04a 0.80 ± 0.02a

5 0.51 ± 0.01bc 0.67 ± 0.04b 0.79 ± 0.02a 0.80 ± 0.01a

10 0.50 ± 0.02bcd 0.63 ± 0.01bc 0.73 ± 0.03bc 0.80 ± 0.03a

25 0.49 ± 0.01cd 0.61 ± 0.02c 0.61 ± 0.01cd 0.79 ± 0.03a

50 0.48 ± 0.03d 0.49 ± 0.06d 0.6 0 ± 0.01cd 0.79 ± 0.03a

100 0.37 ± 0.04e 0.49 ± 0.02d 0.49 ± 0.02e 0.74 ± 0.01bc

200 0.35 ± 0.04e 0.47 ± 0.05d 0.47 ± 0.03e 0.73 ± 0.02bc

300 0.35 ± 0.01e 0.46 ± 0.02d 0.47 ± 0.03e 0.60 ± 0.01d

400 0.34 ± 0.03e 0.46 ± 0.02d 0.47 ± 0.01e 0.60 ± 0.02d

Data represent mean ± standard deviation of 6 replicates. Different letters show statistically difference at p < 0.05 (ANOVA)

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Table 10.4 Effect of illicit drugs on Proline content (µM g-1 fresh weight) in L. minor L. 7d exposure.

Concentration Proline (µM g-1 fw) ( mg L-1) Cocaine MAP MDMA PSE

0 4.7 ± 0.04a 4.7 ± 0.04a 4.7 ± 0.04a 4.7 ± 0.04a

1 4.7 ± 0.07a 4.7 ± 0.12a 4.7 ± 0.10a 4.7 ± 0.04a

5 5.0 ± 0.07ab 4.9 ± 0.09ab 4.7 ± 0.02a 4.7 ± 0.08a

10 5.2 ± 0.14ab 4.9 ± 0.03ab 4.7 ± 0.04a 4.7 ± 0.07a

25 5.3 ± 0.11c 5.0 ± 0.03c 4.8 ± 0.06b 4.7 ± 0.10a

50 5.4 ± 0.08de 5.2 ± 0.17de 4.9 ± 0.17c 4.7 ± 0.11a

100 5.4 ± 0.09de 5.2 ± 0.12de 5.1 ± 0.08d 4.8 ± 0.12b

200 5.4 ± 0.05de 5.2 ± 0.04de 5.1 ± 0.09d 4.8 ± 0.07b

300 5.5 ± 0.14f 5.3 ± 0.09f 5.1 ± 0.17d 4.9 ± 0.09c

400 5.5 ± 0.09f 5.4 ± 0.07g 5.1 ± 0.12d 4.9 ± 0.14c

Data represent mean ± standard deviation of 3 replicates. Different letters show statistically difference at p < 0.05 (ANOVA)

10.4 Conclusion Our results show that all tested growth and biochemical parameters are sensitive and provide information regarding the toxicity of illicit drugs on L. minor L.: cocaine was more toxic than MAP, MDMA and PSE. The highest toxicity of cocaine and least toxicity of PSE to L. minor L. is possibly due to: (i) rapid transformation of cocaine into stable toxic metabolites (BE and EME); and (ii) almost complete disappearance of PSE in 7d exposure. Of all the growth parameters, fresh weight appears to be slightly more sensitive to cocaine, MAP and MDMA, and therefore may be recommended to be used as a biomarker. It is clear that illicit drugs pose a potential risk to aquatic plants, even at smaller concentrations. L. minor L. is very useful for biomonitoring illicit drugs’ contamination in aquatic systems. This is evident from our results where high phytotoxic sensitivity against illicit drugs was documented.

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10.5 References

ACC (Australian Crime Commission), illicit drug data report 2012-13. http://www.crimecommission.gov.au/publications/illicit-drug-data-report/illicit-drug- data-report-2012-13 [Accessed on 10th May 2014]. Alia PP, Saradhi PP. Proline Accumulation under heavy metal stress. J Plant Physiol 1991; 138:554-558. Appenroth KJ, Krech K, Keresztes Á, Fischer W, Koloczek H. Effects of nickel on the chloroplasts of the duckweeds Spirodela polyrhiza and Lemna minor and their possible use in biomonitoring and phytoremediation. Chemosphere 2010; 78:216-223. Ashraf M, Foolad MR. Roles of glycine betaine and proline in improving plant abiotic stress resistance. Environ Exper Bot 2007; 59:206–216. Bates L, Waldren R, Teare I. Rapid determination of free proline for water-stress studies. Plant and soil 1973; 39:205-207. Binelli A, Pedriali A, Riva C, Parolini M. Illicit drugs as new environmental pollutants: cyto- genotoxic effects of cocaine on the biological model Dreissena polymorpha. Chemosphere 2012; 86:906–11. Crist RH, Oberholser K, Wong B, Crist DR. Amine-algae interactions: Cation exchange and possible hydrogen bonding. Environ Sci Technol 1992; 26:1523-1526. Delgado E, Parry MAJ, Vadell J, Lawlor DW, Keys AJ, Medrano H. Effect of water-stress on photosynthesis, leaf characteristics and productivity of field grown Nicotiana tabacum L. genotypes selected for survival at low CO2. J Exp Bot 1992; 43:1001-1008. Drost W, Matzke M, Backhaus T. Heavy metal toxicity to Lemna minor: studies on the time dependence of growth inhibition and the recovery after exposure. Chemosphere 2007; 67:36-43. EMCDDA (European Monitoring Centre for Drugs and Drug Addiction). 2013. Drug profiles. http://www.emcdda.europa.eu/publications/drug-profiles/mdma. [Accessed on October 20, 2013] Eskandari MR, Rahmati M, Khajeamiri AR, Kobarfard F, Noubarani M, Heidari H. A new approach on methamphetamine-induced hepatotoxicity: involvement of mitochondrial dysfunction. Xenobiotica 2013; 44:70-76. Filippou P, Bouchagier P, Skotti E, Fotopoulos V. Proline and reactive oxygen/nitrogen species metabolism is involved in the tolerant response of the invasive plant species Ailanthus altissima to drought and salinity. Environ Exp Bot 2014; 97:1-10. Gubbins EJ, Batty LC, Lead JR. Phytotoxicity of silver nanoparticles to Lemna minor L. Environ Pollut 2011; 159:1551-1559. 173

ISO/DIS 20079, 2004. Water quality – Determination of the toxic effect of water constituents and wastewater on duckweed (Lemna minor) – Duckweed growth inhibition test. ISO 7888, 1985. Lichtenthaler HK. Chlorophylls and carotenoids – pigments of photosynthetic biomembranes. Methods Enzymol 1987; 148:350-82. Mitsou K, Koulianou A, Lambropoulou D, Pappas P, Albanis T, Lekka M. Growth rate effects, responses of antioxidant enzymes and metabolic fate of the herbicide Propanil in the aquatic plant Lemna minor. Chemosphere 2006; 62:275-284. Moser B, Rayburn J. Evaluation of developmental toxicity of interaction between caffeine and pseudoephedrine using frog embryo teratogenesis assay-Xenopus (Fetax). Biosci 2007; 78:1-9. Naumann B, Eberius M, Appenroth KJ. Growth rate based dose-response relationships and EC-values of ten heavy metals using the duckweed growth inhibition test (ISO 20079) with Lemna minor L. clone St. J Plant Physiol 2007; 164:1656-1664. Nedjimi B, Daoud Y. Cadmium accumulation in Atriplex halimus subsp. schweinfurthii and its influence on growth, proline, root hydraulic conductivity and nutrient uptake. Flora - Morpho, Distribu, Functio Ecol Plant 2009; 204:316-324. Organisation for Economic Co-operation and Development. OECD Guidelines for the testing of Chemicals 221. Lemna sp. Growth Inhibition Test, 2006. Paris, France Pal R, Megharaj M, Kirkbride KP, Heinrich T, Naidu R. Illicit drugs and the environment—A review. Sci Total Environ 2013; 463-464:1079-1092. Parolini M, Binelli A. Adverse effects induced by ecgonine methyl ester to the zebra mussel: A comparison with the benzoylecgonine. Environ Pollut 2013; 182:371-378. Parolini M, Pedriali A, Riva C, Binell A. Sub-lethal effects caused by the cocaine metabolite benzoylecgonine to the freshwater mussel (Dreissena polymorpha). Sci Total Environ 2013; 444:43-50. Peraile I, Granado N, Torres E, Gutiérrez-López MD, Moratalla R, Colado MI, O’shea E. Cocaine potentiates MDMA-induced oxidative stress but not dopaminergic neurotoxicity in mice: implications for the pathogenesis of free radical-induced neurodegenerative disorders. Psychopharmacol 2013; 230:125-135. Pomati F, Castiglioni S, Zuccato E, Fanelli R, Rossetti C, Calamari D. Effects of environmental contamination by therapeutic drugs on human embryonic cells. Environ Sci Technol 2006; 40:2442-7. Radić S, Stipaničev D, Cvjetko P, Mikelić IL, Rajčić MM, Širac S, Pevalek-Kozlina B, Pavlica M. Ecotoxicological assessment of industrial effluent using duckweed (Lemna minor L.) as a test organism. Ecotoxicol 2010; 19:216-222.

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Rai V, Vajpayee P, Singh SN, Mehrotra S. Effect of chromium accumulation on photosynthetic pigments, oxidative stress defence system, nitrate reduction, proline level and eugenol content of Ocimum tenuiflorum L. Plant Sci 2004;167:1159-1169. Rosi-Marshall EJ, Snow D, Bartelt-Hunt SL, Paspalof A, Tank JL. A review of ecological effects and environmental fate of illicit drugs in aquatic ecosystems. J Hazard Mater 2015; 282:18-25. Spoljarevic M, Agic D, Lisjak M, Gumze A, Wilson, ID, Hancock JT, Teklic T. The relationship of proline content and metabolism on the productivity of maize plants. Plant Signal Behav 2011; 6:251–257. UNODC (United Nations Office on Drugs and Crime). Amphetamines and Ecstasy: Global ATS Assessment. United Nations Publication; 2008.http://www.unodc.org/documents/scientific/ATS/Global-ATS-Assessment-2008- Web.pdf [Accessed on February 15, 2012]. UNODC (United Nations Office on Drugs and Crime).World Drug Report. United Nations Publication.2014. http://www.unodc.org/unodc/secured/wdr/wdr2014/World_Drug_Report_2014.pdf [Accessed on September 25, 2014]. Voetberg GS, Sharp RE. Growth of the maize primary root at lo0w water potentials. Role of increased proline deposition in osmotic adjustment. Plant Physiol 1991; 96:1125-30. Zezulka Š, Kummerová M, Babula P, Váňová L. Lemna minor exposed to fluoranthene: Growth, biochemical, physiological and histochemical changes. Aqua Toxicol 2013; 140-141:37-47. Zobiole LHS, Kremer RJ, Oliveira JR, Constantin J. Glyphosate affects chlorophyll, nodulation and nutrient accumulation of “second generation” glyphosate-resistant soybean (Glycine max L.). Pest Biochem Physiol 2011; 99:53-60. Zuccato E, Castiglioni S, Bagnati R, Chiabrando C, Grassi P, Fanelli R. Illicit drugs, a novel group of environmental contaminants. Water Res 2008; 42:961-8.

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Chapter 11

Summary and conclusions

Illicit drug abuse is a serious global problem that can lead to significantly damaging effects on human health and social welfare. The major illicit drugs fall into the categories of opiates, cocaine, cannabis, and amphetamine-type stimulants. Globally, several million people are reported to be active users of these illicit compounds. Our environment is the ultimate destination for all these compounds followed by their metabolism in the human body and/or accidental or deliberate disposal of illicit drugs and associated compounds from the illicit drugs’ manufacturers. There are reports on the presence of illicit drugs and their metabolites in various environmental niches from many parts of the world. Although considerable information on the environmental occurrence and mammalian toxicity of these chemicals exists in the literature, in Australia particularly South Australia, information is lacking on their occurrence and distribution, and their environmental fate and ecotoxicity. This study addresses these knowledge gaps of illicit drugs in wastewater (influent and effluent), surface waters, sewage sludge and sediments surrounding the city of Adelaide and their fates in different soils. In addition to this, aquatic (Daphnia carinata and Lemna minor L.) and terrestrial (Eisenia fetida) toxicity of illicit drugs were also investigated.

In this study, the occurrence and distribution pattern of illicit drugs were investigated by analysing the wastewater and sewage sludge collected from three metropolitan WWTPs and surface waters from adjoining areas of Adelaide. MAP, MDMA and BE were the most abundant and ubiquitous compounds in all the environmental samples analysed. The environmental contamination of illicit drugs could have originated from wastewater from WWTPs. Although the concentrations of these contaminants in the environment are quite low, their impact on aquatic organisms and risk to human health cannot be ignored. The ability of soil particles to bind and release MAP and MDMA varied considerably based on the soil type. Moreover, soil factors such as organic carbon, dissolved organic carbon and clay had positively influenced the sorption of MAP and MDMA while cation exchange capacity, electrical conductivity and sand content exerted a negative influence on sorption. Sorption and desorption of the tested illicit drugs in soils followed the order: MAP ˃ MAP mixture ˃ MDMA mixture ˃ MDMA; MAP ˃ MAP mixture ˃ MDMA ˃ MDMA mixture. These outcomes do indicate that sorption and desorption patterns and also provide valuable evidence about the transport, fate and risk of MAP and MDMA in the environment. Degradation of cocaine in soil is a complex process which depends on various factors. 176

Generally, the degradation pattern of cocaine mostly relied both on biotic and some abiotic processes in the soils. BE was relatively stable for a period of time in non-sterile soil compared to cocaine. Cocaine degrades very rapidly in soil into BE and EME which are comparatively more toxic than cocaine on aquatic biota reported in some studies. Hence, more studies are required on adverse effects of cocaine and its metabolites on soil biota.

The acute and genotoxicity test revealed that MAP, MDMA, cocaine and PSE were toxic to aquatic invertebrates such as Daphnia carinata. The overall toxicity of these substances on D. carinata followed the order: cocaine ˃ MAP ˃ MDMA ˃ PSE. However the toxicity of illicit drugs in waters varied considerably due to the influence of various water quality parameters such as suspended particulate matter, DOC, pH, Cl- , Ca & Mg. The acute toxicity of cocaine, MAP, MDMA and PSE in water to D. carinata indicated the following trend: cladoceran water ˃ sterile natural water ˃ non-sterile natural water. In terms of genotoxicity, all these illicit compounds exhibited significant DNA damage and olive tail movement to D. carinata. It is clear that even low level chronic exposure of these compounds to D. carinata cause serious harmful effects including genetic material damage. An aquatic plant (Lemna minor L.) exposed to illicit drugs showed that all growth and biochemical parameters tested were sensitive and provided information on the phytotoxicity of illicit drugs. Cocaine was more toxic to L. minor than MAP, MDMA and PSE in 7d exposure. Results suggested that illicit drugs pose a potential risk to aquatic plants, even in only smaller concentrations. L. minor L. is very useful for biomonitoring illicit drugs’ contamination in aquatic systems. Tests on the toxicity of illicit drugs to earthworms (Eisenia fetida) demonstrated that despite the fact that the illicit drugs did not cause mortality to earthworms, they did compromise their growth, reproduction, morphology, behaviour and caused genetic damage.

11.1 This research has demonstrated that:

It is evident that MAP, MDMA and cocaine were the most commonly abused illicit drugs in the Adelaide region. The primary source of illicit drugs contaminating the environment may arise from the use of WWTPs treated water. Furthermore sorption and desorption of MAP and MDMA in soil were influenced by its physico-chemical properties. MAP has greater sorption and desorption potential compared to MDMA regardless of soil tested. This study also shows that cocaine degradation in soil occurs more rapidly under non-sterile conditions. BE and EME are the main degradation products from cocaine in the soil environment. Cocaine and MAP were more toxic than MDMA and PSE to the aquatic organism Daphnia carinata. The toxicity of illicit drugs in water on D. carinata was iinfluenced by the 177

water quality parameters. Sub-lethal exposure of cocaine, MAP, MDMA and PSE contributed to the severe ill effects in D. carinata including genetic damage. Comet assay in Daphnia could serve as a useful tool for monitoring the genotoxic effects of illicit drugs in water bodies. With reference to L. minor L. all the tested growth and biochemical parameters were sensitive to illicit drugs, and cocaine was more toxic to L. minor L. than MAP, MDMA and PSE. Illicit drugs pose a potential risk to aquatic plants, even at smaller concentrations and L. minor L. appears to be very useful for biomonitoring illicit drugs contamination in aquatic systems. Finally, illicit drugs (cocaine, MAP, MDMA and PSE) affected the growth of Eisenia fetida, its reproduction, morphology and behaviour despite the fact that no mortality occurred when exposed to illicit drugs.

11.2 Propositions for future research:

Although the soils used in this study are representative of many found throughout Australia, the response of other types of soil upon exposure to commonly abused illicit drugs and related precursors, reagents, and manufacturing by-products should be investigated. It is possible that the effects of two chemicals acting together will be greater than the effect of each chemical individually, or the sum of the individual effects. It is therefore important to define the degradation patterns using suitable mixtures of target compounds in aquatic and terrestrial situation, and a parallel toxicity test with various biota. Investigations using real contaminated soils are needed in order to validate biomarkers in illicit drugs contaminated site. More studies are required on the chronic effects at environmental concentrations of commonly abused illicit drugs to soil and aquatic biota.

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