EARTHWORM BIOMARKERS FOR MONITORING PERSISTENT ORGANIC POLLUTANTS

Transcription

1 EARTHWORM BIOMARKERS FOR MONITORING PERSISTENT ORGANIC POLLUTANTS SRINITHI MAYILSWAMI B.Sc. (Agriculture), TNAU M.Sc. (Environmental Sciences), UEA A thesis presented in partial fulfilment of the requirements for the degree of Doctor of Philosophy Centre for Environmental Risk Assessment and Remediation Division of the Information Technology, Engineering and the Environment August 2014

2 TABLE OF CONTENTS TABLE OF CONTENTS... 2 LIST OF FIGURES... 9 LIST OF TABLES LIST OF ABBREVIATIONS ABSTRACT DECLARATION ACKNOWLEDGEMENT Introduction Persistent organic Pollutants Significance of earthworm in ecotoxicological studies Earthworm biomarkers Literature Review POPs (Persistent Organic Pollutants) Earthworm Biomarkers Biomarkers in earthworms Role of antioxidant enzymes in defence mechanism Detoxification of organics by biotransformation of enzymes Biomarkers at cellular level Genotoxic biomarkers Biomarkers on blood chemistry Behavioural response biomarkers Impairment in reproduction Molecular Biomarkers Conclusion Material and Methods Soils and earthworms Earthworm exposure Cellulase activity Collection of coelomocytes Cell viability test

3 3.6 Comet assay Total Antioxidant capacity Lipid peroxidation Neutral red retention Assay MTT assay LDH cytotoxicity Calculation of PFOA and PFOS in soil and animal tissue Extraction of PAHs Avoidance test Locomotion Cast production Wound healing capacity Statistics Gene expression study Chemicals Animal treatment Experimental Design RNA Isolation and next generation sequencing Sequence analysis Acute toxicity of PFOS Introduction Materials and Methods Results Survival and growth of earthworms Lysosome membrane stability Cellulase activity Avoidance behavior Reproduction Discussion Mortality and Weight loss Cellulase activity Neutral red retention assay Avoidance behaviour Reproduction rate Conclusion

4 5 Acute toxicity of PFOA Introduction Materials and Methods: Results Survival and growth Lysosome membrane stability Cellulase activity Avoidance behaviour Reproduction data Discussion Mortality and Weight change Cellulase activity Effect on lysosome membrane stability Avoidance in worms Reproduction effect Conclusion Genotoxicity of PFOS and PFOA Introduction Materials and Methods Results PFOS induced DNA damage PFOA induced DNA damage Total antioxidant capacity Lipid peroxidation Effect of PFOS and PFOA on E.fetida defense system Lipid peroxidation induced by PFOS and PFOA in earthworms DNA damage induced by PFOS and PFOA Conclusion Accumulation of PFOS Introduction Materials and Methods Results Bioaccumulation of PFOS Cytotoxicity of PFOS Locomotion

5 7.3.4 Wound healing Capacity Cast production Discussion Bioaccumulation of PFOS Effect on cell viability Effect on Locomotion ability Effect on wound healing capacity Effect on Cast production Conclusion Accumulation of PFOA Introduction Materials and methods: Results Bioaccumulation of PFOA in earthworms Cytotoxicity of PFOA Physiological end points Discussion Bioaccumulation of PFOA in earthworm Effect on cell viability Behavioral responses Effect of PFOA on locomotion Effect on Cast production Effect of PFOA on the wound healing capacity Conclusion Application of biomarker battery for evaluation of Benzo(a)Pyrence effects on earthworm Introduction Materials and Methods Results Acute toxicities highlighting survival, weight loss, cellulase activity and bioaccumulation Cytotoxicity assay and membrane stability Lipid peroxidation, antioxidant capacity and DNA damage Behavioral response to B(a)P Reproduction effects Discussion

6 9.4.1 Survival, growth, cellulase activity and accumulation of B(a)P Biochemical responses and genetic damages Membrane stability and cytotoxicity Behavioral response Conclusion Effect of phenanthrene on the biological responses of earthworm under laboratory conditon Introduction Material and Methods Results Acute toxicities highlighting survival, weight loss, cellulase activity and bioaccumulation Cytotoxicity assay and membrane stability Lipid peroxidation, antioxidant capacity and DNA damage Behavioral response to phenanthrene Reproduction effects Discussion Survival, weight loss, cellulase, bioaccumulation and reproduction Lysosome membrane stability, cell cytotoxicity Antioxidant capacity, lipid peroxidation and DNA damage Behavioral responses Conclusion Differential response of biomarkers in the body of earthworm exposed to pyrene under laboratory condition Introduction Materials and methods Results Accumulation, survival, growth and cellulase activity Cytotoxicity assay and membrane stability Lipid peroxidation, antioxidant capacity and DNA damage Behavioural response to phenanthrene Reproduction effect induced by pyrene Discussion Accumulation, survival, growth, cellulase activity Antioxidant capacity, lipid peroxidation and oxidative DNA damage NRRT, MTT and LDH cytotoxicity

7 Behavioural response Conclusion Gene expression study on earthworms exposed to PFOS Introduction Materials and Methods Results and Discussion Transcript assembly and annotation Differential gene expression Up regulated genes involved in calcium signaling Down regulated genes involved in calcium signaling Oxidative damage response genes Fertility related genes Gene involved in apoptosis PFOS toxicity Neuronal damages caused by PFOS Over expressed genes that are involved in neuronal development Down regulated genes involved in neuronal development Conclusion Gene expression changes in earthworm when chronically exposed to PFOA Introduction Materials and Methods Results and Discussion Transcript assembly and annotation Differential gene expression Calcium homeostasis and related genes Neuronal development related genes Reproduction related genes Lipid metabolism and modification Genes involved in apoptotic process Ungrouped genes Conclusion Differential gene expression in earthworm when chronically exposed to Benzo(a)Pyrene Introduction Materials and Methods

8 14.3 Results and Discussion Transcript assembly and annotation Differential gene expression Calcium binding and homeostasis Apoptotic process related genes Gene that are involved in cell projection and cytoskeleton Protein localization and transport related genes Nucleotide binding proteins Other genes and functions Conclusion Transcriptonme anlaysis in earthworm chronically exposed to pyrene Introduction Materials and methods Results and discussion Transcript assembly and annotation Differential gene expression Cell projection organization and intrinsic to membrane Cytoskeletal genes Protein catabolic process Nucleotide binding proteins Ungrouped Conclusion Summary and Conclusions Contribution to science Recommendation of future research References

9 LIST OF FIGURES Figure 2.1 Earthworm biomarkers Figure 3.1 Pictorial representation of earthworm avoidance test procedure Figure 3.2 Locomotion pattern exhibited by both control and treatment worms Figure 4.1 Survival percentage of earthworms treated with PFOS Figure 4.2 Weight loss percentage in earthworms exposed to PFOS Figure 4.3 Lysosome membrane stability in earthworms treated with PFOS 69 Figure 4.4 Cellulase activity in earthworms exposed to PFOS Figure 4.5 Avoidance pattern exhibited by earthworms exposed to PFOS 71 Figure 5.1 Survival percentage of earthworms treated with PFOA Figure 5.2 Weight loss percentage in earthworms exposed to PFOA Figure 5.3 Lysosome membrane stability in earthworms treated with PFOA 83 Figure 5.4 Cellulase activity in earthworms exposed to PFOA Figure 5.5 Avoidance pattern exhibited by earthworms exposed to PFOA 84 Figure 6.1 Comet assay image showing DNA damage in earthworms exposed to PFOS and PFOA.. Figure 6.2 PFOS induced DNA damage 95 Figure 6.3 PFOA induced DNA damage 97 Figure 6.4 Total antioxidant capacity in earthworms exposed to PFOS and PFOA Figure 6.5 Lipid peroxidation in earthworms exposed to PFOS and PFOA Figure 7.1 Bioaccumulation factor of PFOS in earthworms Figure 7.2 Cytotoxicity percentage observed in earthworms treated with PFOS Figure 7.3 Figure 7.4 Mean crawled distance of earthworms exposed to different concentration of PFOS..... Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure to PFOS. Figure 7.5 Effect of PFOS on earthworms cast production 114 Figure 8.1 Bioaccumulation factor of PFOA in earthworms Figure 8.2 Cytotoxicity percentage observed in earthworms treated with PFOA Figure 8.3 Figure 8.4 Figure 9.1 Figure 9.2 Mean crawled distance and cast production of earthworms exposed to different concentration of PFOA... Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure to PFOA... Effect of Benzo(a)Pyrene on earthworms (a) survival (b) weight loss (c) bioaccumulation and (d) cellulase activity Cytotoxicity percentage observed in earthworms treated with Benzo(a)Pyrene. 139 Figure 9.3 Lysosome membrane stability in earthworms exposed to B(a)P Figure 9.4 Effect of B(a)P on earthworms antioxidant capacity

10 Figure 9.5 DNA damage induced by B(a)P in earthworms 141 Figure 9.6 Effect of B(a)P on earthworms lipid peroxidation 142 Figure 9.7 Figure 9.8 Figure 10.1 Figure 10.2 Figure 10.3 (a) Mean crawled distance by earthworm in one minute on 14 day and cast production (g/worm/day) on day 14 in earthworms exposed to B(a)P in OECD and Neutral soils (b) Avoidance behaviour of earthworms exposed to B(a)P (c) and (d) Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure. Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure in alkaline soils Effect of phenanthrene on earthworms (a) survival (b) weight loss (c) bioaccumulation and (d) cellulase activity Cytotoxicity percentage observed in earthworms treated with phenanthrene. 160 Lysosome membrane stability in earthworms exposed to phenanthrene. Figure 10.4 Effect of phenanthrene on earthworms antioxidant capacity 161 Figure 10.5 DNA damage induced by phenanthrene in earthworms 162 Figure 10.6 Effect of phenanthrene on earthworms lipid peroxidation 162 Figure 10.7 Figure 11.1 (a) Mean crawled distance by earthworm in one minute on 14 day and cast production (g/worm/day) on day 14 in earthworms exposed to phenanthrene in OECD and Neutral soils (b) Avoidance behaviour of earthworms exposed to phenanthrene (c) and (d) Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure. Effect of pyrene on earthworms (a) survival (b) weight loss (c) bioaccumulation and (d) cellulase activity Figure 11.2 Cytotoxicity percentage observed in earthworms treated with pyrene 180 Figure 11.3 Lysosome membrane stability in earthworms exposed to pyrene 182 Figure 11.4 Effect of pyrene on earthworms antioxidant capacity Figure 11.5 DNA damage induced by pyrene in earthworms Figure 11.6 Effect of pyrene on earthworms lipid peroxidation Figure 11.7 Figure 11.8 (a) Mean crawled distance by earthworm in one minute on 14 day and cast production (g/worm/day) on day 14 in earthworms exposed to pyrene in OECD and Neutral soils (b) Avoidance behaviour of earthworms exposed to pyrene (c) and (d) Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure. Percent of earthworms Eisenia fetida having complete healing 5 d post wounding after 7, 14 and 28-d exposure in alkaline soils

11 Table 2.1 LIST OF TABLES Priority persistent organic pollutant targeted in Stockholm convention... Table 2.2 Detailed table on the history of POPs existence in the environment 25 Table 2.3 Important biomarkers in earthworms exposed to various organic contaminants.. 33 Table 2.4 Biomarkers of cellular level in earthworms Table 3.1 Different soil properties of OECD alkaline and neutral soils Table 4.1 LC50 value of PFOS on earthworms in three different soil types. 69 Table 4.2 Cocoon production and juvenile emergence in earthworms on 28 th and 56 th day exposed to PFOS. 71 Table 5.1 LC50 value of PFOA on earthworms in three different soil types 82 Table 5.2 Cocoon production and juvenile emergence in earthworms on 28 th and 56 th day exposed to PFOA 85 Table 9.1 Cocoon production and juvenile emergence in earthworms on 28 th and 56 th day exposed to B(a)P. 146 Table 10.1 Cocoon production and juvenile emergence in earthworms on 28 th and 56 th day exposed to phenanthrene. 167 Table 11.1 Cocoon production and juvenile emergence in earthworms on 28 th and 56 th day exposed to pyrene 187 Table 12.1 Genes that are up-regulated in response to chronic PFOS exposure to earthworms 205 Table 12.2 Genes that are down-regulated in response to chronic PFOS exposure to earthworms Table 13.1 In Eisenia fetida genes altered by PFOA that are involved in calcium homeostasis 212 Table 13.2 In Eisenia fetida genes altered by PFOA that are involved in reproduction. 213 Table 13.3 In Eisenia fetida genes altered by PFOA that are involved in neural development Table 13.4 Table 13.5 Table 13.5 Table 14.1 Table 14.2 Table 14.3 Table 14.3 In Eisenia fetida genes altered by PFOA that are involved in lipid metabolism. 215 In Eisenia fetida genes altered by PFOA that are involved in apoptotic process In Eisenia fetida genes altered by PFOA that are ungrouped Calcium ion binding protein transcripts that are altered by B(a)P in earthworms Gene transcripts that are altered by B(a)P involved in apoptotic process Gene transcripts that are altered by B(a)P involved in cell projection and cytoskeletio 227 Gene transcripts that are altered by B(a)P involved in protein localisation and transport

12 Table 14.4 Ungrouped genes that are altered by B(a)P exposure Table 15.1 Transcripts that are altered by pyrene involved in cell projection and organisation 237 Table 15.2 Genes that are altered by pyrene involved in cytoskeletal structure 238 Table 15.3 Genes that are altered by pyrene involved in protein catabolic processes 238 Table 15.4 Genes that are altered by pyrene involved in nucleotide binding property 238 Table 15.4 Ungrouped genes that are altered by pyrene exposure

13 LIST OF ABBREVIATIONS BAF PFOS PFOA GST BSAF BMF B(a)P NRR NRRT SOD LCMS HPLC Bioaccumulation factor Perflourooctanesulfonic acid Perfluorooctanoic acid Glutathione-S-transferase Biota sediment accumulation factor Biomagnification factor Benzo(a)Pyrene Neutral red retention Neutral red retention time Super oxide dismutase Liquid chromatography mass spectroscopy High performance liquid chromatography 13

14 ABSTRACT Historically, public health regulations have been based on theoretical risk calculations according to known levels of chemical substances in air, water, soil, food, other consumer products and sources of potential exposure. Biomonitoring offers the opportunity to analyse the actual internal levels of bodily substances from all potential routes of exposure at one time, which may contribute to improving risk assessments. With regard to ecological risk assessment, biomarkers allow exposure of pollutants to be detected and populations at risk to be identified. It is expected that in time, biomarkers will be used as a tool in bio-monitoring soon form the foundation for ecosystem health tracking system and includes identification of environmental sources, exposure and population effects. Persistent organic pollutants, or POPs, are organic pollutants that are resistant to environmental degradation by biological, physical and chemical processes. They also have great tendency to travel long distances, bioaccumulate in human tissues and biomagnify in food chain; and through this they impose serious threat to human and ecosystem health. My research project examines the relationships between biochemical and physiological biomarkers (survival, growth, cellulase activity, lysosome membrane stability, antioxidant capacity, lipid peroxidation, cellular damage, DNA damage, behavioural response) and accumulated concentrations of contaminants such as perfluoro chemicals (PFOS, PFOA) and polycyclic aromatic hydrocarbons i(b(a)p, phenanthrene and pyrene) in the keystone species Eisenia fetida. Hence the purpose of this research is to couple ecotoxicity studies through bioassays and chemical analysis. The other part of this thesis examined the gene expression pattern of earthworms after chronic exposure to PFOS, PFOA and PAHs at low level concentrations. In the toxicity analysis, the impact of PFOS and PFOA in different soils was evaluated by integrating the earthworm bioassay. The LC 50 value for PFOS, based on the measured concentration rather than the nominal concentration used for spiking in the present study were ±3.9 mg PFOS kg -1 in alkaline soil, 446.8±3.3 mg PFOS kg -1 in neutral soil and 159.9±2.8 mg PFOS kg -1 in OECD soil. Similarly LC 50 value for PFOA were 823.8±5.3 mg PFOA kg -1 in alkaline soil, 894.9±5.9 mg PFOA kg -1 in neutral soil and 672.1±7.86 mg PFOA kg -1 in OECD soil. Ecotoxicity studies with the PFOS and PFOA demonstrated that PFOS was more toxic to earthworms than PFOA. Even though the toxicity of PFOA is low they are capable of inducing significant reproductive and genetic damages in earthworms. This integrated approach clearly defines the ecological impact of perfluorinated contaminants. 14

15 Bioaccumulation study on both PFOS and PFOA concluded that perflourinated chemicals are highly bioaccumulative in earthworm tissues. The bioaccumulation factor was higher at lower concentration exposure. Similarly OECD soil facilitated the negative impacts of both chemicals compared to natural soil. Ecologically relevant parameters like locomotion, cast production and wound healing capacity were highly susceptible to PFOS and PFOA exposure. Also the concentrations below LC 50 value were capable of inducing significant cellular damages in earthworms. Toxicity study on PAHs revealed that B(a)P was highly toxic compared to phenanthrene and pyrene. The LC 50 values ranged from 73 to 150 mg B(a)P kg -1, 220 to 280 mg phenanthrene kg -1 and 520 to 1201 mg pyrene kg -1 in alkaline, neutral and OECD soil. Here also OECD soil showed higher toxicity than natural soils. The order of genotoxicity in earthworms were pyrene < phenanthrene < B(a)P. Bioaccumulation study showed that B(a)P was highly accumulative than phenanthrene and pyrene but low molecular weight compounds also gets accumulated in earthworm tissue to a lesser extent. Depleted levels of ecological parameters could be a consequence of combating toxicants predicted at higher level of biological organisation which was only significant for B(a)P exposure not to phenanthrene and pyrene. The genesis of fresh technologies from the genomics revolution will change the possible utilization of biomarkers to evaluate how pollutants affect people, animals, and ecosystems. Genetic databases supply wide information from which suitable molecular responses can be recognized and, later, utilized to point these problems. From our study we derived the link between novel genetic expression and ecologically relevant whole-organism life-cycle traits. Chronic exposure of Eisenia fetida to PFOS alters the expression of calcium homeostasis related genes and neuronal development related genes whereas PFOA affected the genes related to reproductive system of earthworms. Keywords Earthworms, Biomarkers, Reproduction, Bioaccumulation 15

16 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. Srinithi Mayilswami Signed Date 16

17 ACKNOWLEDGEMENT First and foremost I offer my sincerest gratitude to my supervisor, Prof Megharaj Mallavarapu, who has supported me throughout my thesis with his patience and knowledge whilst allowing me the room to work in my own way. One simply could not wish for a better or friendlier supervisor. During more than a decade of knowing him, he has been instrumental to me for seeing life and science in their full depth, and taught me how to appreciate the good scientific work that helps other researchers to build on it. It has been a privilege to have you as my supervisor and I will always bear in mind your enthusiasm and skilful guidance. I would like to convey my heartfelt thanks and sincere gratitude to my project co-supervisors Dr Kannan Krishnan and Prof. Ravi Naidu for their excellent direction and valuable guidance during all stages of my candidature. Looking back, I am surprised and at the same time very grateful for all I have received throughout these years. It has certainly shaped me as a person and has led me where I am now. All these years of PhD studies are full of such gifts. The main part of this thesis has been conducted at the CERAR, University of South Australia, during the year under the project name Earthworm biomarkers for monitoring persistent organic pollutants. The project has been supported by the Australian government grant (IPRS) and Cooperative Research Centre for Contamination Assessment and Remediation of the Environment (CRC CARE). A large number of people at the centre have been involved in the work leading to this thesis. I direct my sincere gratefulness to all the people at CERAR for the excellent social atmosphere you have maintained in our years together. I acknowledge the Ramaciotti Centre for Genomics, The University of New South Wales, Sydney, for mrna sequencing and eresearch SA for computing facility. Now, when I am about to complete this thesis, I would like to thank my friends and family members especially my beloved parents and my sister for all support and comfort given me and at last I specially thank Muthurathinavel, dear husband who have been waiting with patience for this day. Finally, I thank Kavitha and Karthik for supporting me throughout all my studies at University, shifting my vast collections of ``stuff'' across most of Mawsonlakes and for providing a home in which to complete my writing up. I finish with a final silence of gratitude for my life. 17

18 1 Introduction 1.1 Persistent organic Pollutants Persistent organic pollutants (POPs) include a manifold class of organic chemicals capable of causing lethal effects, also highlighted for persistent, bio-accumulative and ability to travel long distances. They possess significant physical-chemical properties, which prompt their environmental responsibilities (Wania 2003, 2006). It s necessary to identify their sources and pathways of emission into the environment which are diverse in nature. In terms of law to govern the emerging POPs, is poorly defined due to lack of quantitative aspect with regards to their global source and emission. Predominantly POPs travel via atmosphere but water also responsible for significant transport causing arctic contamination (Lohmann et al. 2007). The Stockholm Convention on POPs is a global treaty (UNEP 2001) that presently govern POPs to protect the human and environmental health. They classified initially POPs under 3 categories as follows: 1. Pesticides: aldrin, chlordane, DDT, dieldrin, endrin, heptachlor, hexachlorobenzene, mirex, toxaphene; 2. Industrial chemicals: hexachlorobenzene, polychlorinated biphenyls (PCBs); and 3. By-products: hexachlorobenzene; polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans (PCDD/PCDF), and PCBs. Aarhus protocol only acknowledged PAHs are POPs (UNECE, 1998). Nevertheless, currently many researchers have pointed that these compounds are also detected during the process of combustion. Importantly rising emerging contaminants in existence were left underestimated in the environment (Muir, DCG & Howard 2006). New screening level risk assessment model has been created to highlight the ways to estimate and prioritize POPs by identifying their environmental behaviours along with their transport, bioaccumulation, and humans and wildlife contamination (Arnot et al. 2006). 18

19 1.2 Significance of earthworm in ecotoxicological studies Growing burden imposed by chemical contamination of soil drags the attention of scientific community and international agencies towards this to focus on controlling their source and emission. The conventional way to estimate soil contamination includes analysis of the concentrations of contaminant in the soil to correlate with specific threshold values, fails to predict the havoc faced by biota specific to contamination. Several crucial conditions have been ignored, mainly toxicity of chemicals not checked in the choice of contaminants to be estimated, also collective responses (synergism and antagonisms) of pollutants on biota and bioavailability are not presented. Bioavailability cite the fraction of pollutant that is absorbed by an organism from the environmental portions (i.e., through both passive and active ways), and causes the toxicity directly (Smith, Reed & Webb 1993). Physiochemical properties of soil fluctuates the bioavailable fraction of contaminants to soil dwelling organism and these include ph, cation exchange capacity, and organic matter content (Bradham et al. 2006; Spurgeon, David J et al. 2006; Criel et al. 2008). This addresses the difficulty in assessing the pollutant bioavailability based on the total concentration and soil characteristics. So in order to bridge the gap we should rely on the exposed organism. This leads to the innovation of new biological way of soil monitoring such as the estimation of biochemical and cellular responses to contaminants (i.e. biomarkers) on sentinel organism (bioindicators), have emerged as significant method to predict the quality of the soil (Kammenga et al. 2000). Bioindicators render much data in relevance to the contaminated sites that may be helpful, yet the productivity from biological techniques relies on the selection of sensitive organism and ecologically oriented parameter observations (Cortet et al. 1999). The earthworms carry important responsibility in decomposition of organic matter and succeeding nutrient management has conceived the idea of earthworms as sentinel organism to assess the biological response induced by soil pollutants. Ultimately earthworms occupy a strong position in ecotoxicology (Spurgeon, David J, Weeks & Van Gestel 2003). Added advantages such as relatively simple to maintain, also aid us to estimate vital life-cycle parameters such as growth and reproduction with respect to accumulation and excretion of contaminants, and biochemical responses. Thus, earthworms are significant in ecotoxicological research. 19

20 1.3 Earthworm biomarkers A biomarker is referred as any alteration in biochemical, cellular, physiological or behavior in tissue or body fluid samples, or at the level of whole organisms that can be detectable to declare the information on exposure and/or effects from one or more pollutants (Depledge 1994). The impacts of pollutants at lower levels of biological organization (e.g. biochemical, cellular, physiological) are commonly detectable as early signs rather than those occurring at higher levels (e.g., ecological effects) and hence they act as prognostic signal for toxicological impacts within populations. Potentially, any alterations in any of the molecular, cellular, biochemical, and physiological responses existing within an organism with respect to pollutant exposure could be applied as biomarkers. Biomarkers are broadly divided in to two sections (1) biomarkers of exposure, and (2) biomarkers of effect (Chambers et al. 2002). The initial one represents that an organism has contact to a contaminant, and provide forewarning sign for exposure to stress condition. They offer qualitative and quantitative measure of vulnerability to diverse pollutants. Nevertheless, the variation in biological response may not be truly lethal impact either on the organism or in the population. Biomarkers of effect are related exactly with the contaminant s mode of action and are scrutinized to link the biological response to the antagonistic property of the contaminants. Overall, application of earthworm biomarker is growing in the field of estimating the effects of contaminants on terrestrial organisms. Predominantly used biomarkers are acetylcholinesterase, metallothionein, biotransformation enzymes and antioxidant defenses because they play vital part in neurocholinergic transmission and in cell homeostasis protecting lethal effects of chemicals (Sanchez-Hernandez, JC 2006; Novais et al. 2011). However, the discovery of novel biomarkers in earthworms is gaining momentum for its efficiency in soil pollution monitoring and survey. In the following Chapter 2, provides a detailed literature review, including a discussion on earthworm biomarkers specific to POPs. This chapter deals specifically with biomarker studies of POPs in earthworms and the potential benefits of using them in environmental assessment. Chapter 3 discusses the materials and methods used throughout the experiment Chapter 4 elaborates on comparative acute and chronic terrestrial ecosystem hazard assessment of PFOS using earthworms (Eisenia fetida) and its effects in 3 different soil conditions. 20

21 Chapter 5 elaborates on comparative acute and chronic terrestrial ecosystem hazard assessment of PFOA using earthworms (Eisenia fetida) and its effects in 3 different soil conditions. Chapter 6 deals with genotoxicity assessment of PFOS and PFOA using three different soils in earthworms (Eisenia fetida) using comet assay. Explanation based on lipid peroxidation and antioxidant capacity has been explained with respect to PFOS and PFOA. Chapter 7 explains the ecological relevant toxicity study of PFOS on earthworm under 3 different soil conditions (Alkaline, OECD and Neutral soil). Bioaccumulation of PFOS in earthworms was also focused in this chapter. Chapter 8 deals with linking different biomarkers in earthworms exposed to PFOA to identify the impact in terrestrial ecosystem with 3 different soil conditions. Special focus was shown on the bioaccumulation of PFOA in earthworm. Chapter 9 Acute toxicity of B(a)P in earthworms maintained in OECD, neutral and alkaline soils. Chapter 10 Acute toxicity of phenanthrene in earthworms maintained in OECD, neutral and alkaline soils. Chapter 11 Acute toxicity of pyrene in earthworms maintained in OECD, neutral and alkaline soils. Chapters 12 to 15 Explores the gene expression pattern in earthworms chronically exposed to PFOS, PFOA, B(a)P and pyrene. Also, a detailed explanation has been dealt with respect to the expression of genes. Chapter 16 summarizes the major findings of this study and suggests future research required in this area. 21

22 2 Literature review 2.1 POPs (Persistent Organic Pollutants) Extensive human activities have resulted in the overwhelming generation of persistent organic pollutants (POPs). These POPs are organic chemicals that can cause much damage to the environment and they are ubiquitous in many terrestrial to ocean ecosystems. They have been detected in the Arctic region and deep ocean depths. POPs can be classified into two types: (1) Natural organic chemicals (2) New organic chemicals Fossil fuel hydrocarbons utilised by humans synthesised and released into the and later discharged into the environment. environment as a result of urbanisation, for example PCBs These organic pollutants are generated either naturally or by human activities, and with their extreme physical and chemical properties these compounds persist in the environment for prolonged duration once discharged. They can remain untouched by physical, chemical and biological degradation mechanisms (Buccini 2003). The reason for this is mainly because of the stability of the carbon and halogen bond which is highly resistant to breakdown. Even though production of certain POPs (DDT, chlorinated pesticides) are strictly prohibited in many countries, they are still present everywhere. The United Nations Environmental Agency implemented strict rules to eliminate the production of these compounds. Slowly these compounds pass from various environmental media into the food chain, and finally enter the human body (Kumar et al. 2005; Zhao, G et al. 2006). Daily human needs are increasing, leading to the invention of new compounds, where POPs like PCBs and perfluorinated chemicals (as in fire-fighting foams) have noxious properties and are dangerous to people s health. Each year new chemicals are being added to the Stockholm Convention list which lists dangerous chemicals (Farrington & Takada 2014). 22

23 Initially 12 chemicals were deemed to be priority contaminants, and later 9 chemicals were added to the Stockholm Convention list. Annex A (Elimination ) Annex B Annex C(Unintentional (Restriction) Production ) Aldrin * DDT * Polychlorinated dibenzofurans (PCDF) Dieldrin * Perfluotooctane Hexachlorobenzene (HCB) sulfonic acid, its salts and Perfluorooctane sulfonyl fluoride # Hexabromobiphenyl # Pentachlorobenzene Alpha Hexachlorocyclohexane # Polychlorinated biphenyls (PCB) Mirex * Technical Endosulfan and Isomers * Chlordane * Endrin * Hexabromodiphenyl ether and heptabromodiphenyl ether # Beta hexachlorocyclohexane * Pentachlorobenzene * # Tetrabromodiphenyl ether and pentabromodiphenyl ether # Chlordecone * Heptachlor * Lindane * Ploychlorinated biphenyle # Toxaphene * Alpha hexachlorocyclohexane 23

24 Beta hexachlorocyclohexane Chlordecone # Hexabromobiphenyl # Hexabromobiphenyl ether and heptabromodiphenyl ether (Commercial octabromodiphenyl ether # Lindane * Pentachlorobenzene * # Perfluorooctane sulfonic acid and perfluorooctane sulfonyl fluoride. # Technical endosulfan and its related isomers * Tetrabromodiphenyl ether and pentabromodiphenyl ether (commercial pentabromodiphenyl ether) # Table 2.1 Priority persistent organic pollutants targeted in Stockholm convention *- By products # - Industrial chemicals ( 24

25 Years POPs Scientific information on distribution of POPS Synthetic organic chemicals World War 2 promoted the production of these chemicals to improve agricultural income, to protect human health and enhance industrialization DDT Effective against pests, so believed to be a boon to agriculture 1945 PFOS 3M company initiated the production of these compounds and production rate increased subsequently later DDT and Pesticides Apart from beneficial effects, many harmful effects associated with the usage of these chemicals have been identified. Notably: Birds fail to reproduce, Population decline in highly polluted areas. 25 Reference (Krueger & Selin 2002; Selin & Eckley 2003) (Krueger & Selin 2002; Selin & Eckley 2003) (Paul, Jones & Sweetman 2008) (Kumar et al. 2005)

26 1970 PFOS Production was around <500 tonnes POPs More faced by the usage of these chemicals, POPs grabs the attention of all the people during this time. 1970s DDT, Rate of application of pesticide Hexacholorocyclohexane, increased to safeguard crops HCB against pests and diseases 1980s POPs Scientists came out with surprising information on nature of these chemicals, their ability to travel longer distances with ease. Research findings highly targeted Artic regions POPs Comprehensive report on release of hazardous POPs from Canada Pesticides Chinese intensive agriculture includes excessive usage of pesticides and insecticides. (Paul, Jones & Sweetman 2008) (Kumar et al. 2005) (Krueger & Selin 2002; Selin & Eckley 2003; Villa et al. 2003; Kumar et al. 2005) (Selin & Eckley 2003) (Selin & Eckley 2003) (El-Shahawi et al. 2010) 26

27 Survey indicated the range of usage Kg/hm 2, which increased around 41.8% annually Toxaphene Reports indicated severe contamination of POPs in largest lake trout in Bow Lake. (Hung et al. 2006) (Also indicated that humans are at high risk subjected to eating the contaminated trout) 1990 POPs (Multiple form) and Taiwan: Danshui river is now PCBs polluted by liquid discharge and atmospheric POPs emitted from disposal of municipal and industrial waste So called Dioxin Belgium scandal. Various food samples have been contaminated by POPs as per the report produced by Polish Veterinary Inspectorate. (Storelli et al. 2004) (Buckley-Golder et al. 1999) 27

28 PFOS Production was at their peak, around 4500 tonnes PCDDs, PCDFs,PCBs, DDE, Marine species from Adriatic sea Organo chlorine pesticides are highly poisoned by organo chemicals which lead to high threat to aquatic ecosystem. Highest level of contamination was noted in Mackerel, Red mullet and Anchovy PFOS 3M company announced their termination of production of PFOS, so there was a sharp reduction in the amount of production, was around 1000 tonnes (Paul, Jones & Sweetman 2008) (Bayarri et al. 2001) (Paul, Jones & Sweetman 2008) Table 2.2 Detailed table on the history of POPs existence in environment 28

29 2.2 Earthworm A flourishing biodiversity is responsible for a well-developed ecosystem and its function. The most common ecosystem engineers are earthworms which inhabit much of the soil ecosystem. Earthworms improve the soil structure and its functions (Paoletti 1999; Jongmans et al. 2003). They are the most popular bioindicators of soil contamination (Cortet et al. 1999; Lanno et al. 2004). Their various activities like feeding, burrowing, and excretion improve soil quality more specifically by aiding water infiltration, facilitating soil aeration and accumulation of organic matter at the top (Morgan et al. 2004). All these activities enable them to have a close contact to the available contaminants especially during ingestion and skin contact (Saxe et al. 2001; Jager et al. 2003; Vijver et al. 2005). Earthworm activity such as modifying the physico-chemical and biological properties of soil act as building blocks for soil formation and soil profile development (Lavelle & Spain 2001). Furthermore, the casts they produce enhance the soil microbial population, indirectly support the flow of nutrients into the soil (Emmerling & Paulsch 2001), and improve the movement of water and air in the soil. Earthworms maintenance of the ecosystem led to the development of the term Earthworm Ecotoxicology (Spurgeon, David J, Weeks & Van Gestel 2003). Being a simple organism with a precise life cycle, we can measure the various lifecycle parameters such as body mass, growth, reproduction factors, bioaccumulation data and various biological responses under controlled conditions without any difficulty. Earthworms are highly preferable for conducting ecotoxicological research (Lionetto, MG, Calisi & Schettino 2012). Interaction with the soil contaminants occurs through various pathways, predominantly through dermal contact and less so through ingestion. Dermal intake generally depends on the concentration of the contaminant in the soil pore water, because earthworms skin is highly porous (Wallwork 1983; Jager et al. 2003; Vijver et al. 2003). Secondly, earthworms feed on the contaminated soil for their survival, ultimately resulting in the ingestion of contaminants through their digestive tract (Morgan et al. 2004). Dermal intake plays a stronger role in contaminant accumulation than ingestion as shown by studies conducted on metal interaction with Lumbricus rubellus (Vijver et al. 2003). 29

30 Pollutants Sensory perception Biochemical changes / Molecular Responses Enzyme induction /Inhibition Mutant gene products New protein formation (Metallothions, stress proteins) Structural alteration of genetic material Physiological responses Metabolism Hormonal imbalance Locomotor activity Water/mineral imbalance Cytological / Morphological responses Lysosome membrane stability Histopathal ogical lesions Gross indices Avoid ance Behavioural changes Mating Feeding Population/community effects Figure 2.1 Earthworm biomarkers Surface migration /Feeding 30

31 2.3 Biomarkers Ecotoxicology deals with the interconnection between toxicology and ecology in a detailed way. Normally contaminants are generated either naturally (metals) or by human involvement (PAHs), and the behavioursr of these chemicals in the ecosystem are examined using ecotoxicology. Ecotoxicology explains the occurrence of contaminants and their impact on biological features of living organisms. As an applied science, it enables ecotoxicologists to create various tools to measure the effect of contaminants well in advance to protect the ecosystem from various risks imposed by these contaminants. Urbanisation leads to contamination in an uncontrolled manner and threatens society through disease, which necessitates more advanced techniques to monitor risks. Biomarkers constitute one such tool that describes the impacts of the contaminants in the biological system more appropriately. Traditional approaches concentrate only on the soil physico-chemical properties to predict the behaviour of contaminants in the soil. Yet biomarkers focus on the bioavailable fraction of the contaminants that pose a threat to living organisms. The biomarker concept must be integrated into an ecological risk assessment to better understand the fate of contaminants (Pauwels et al. 2013). Hypothetically, biomarkers are any alterations that emerged in biological organisations in the body, from the molecular level to behavioural patterns that are measurable. Once an organism is exposed to unwanted conditions there occurs a shift in their health status which can be detected using biomarkers; these are more practical for biomonitoring programs. The general idea behind the term biomarkers is that an organism living in a stressful state either avoids that or suffers, thus the link between stress and suffering forms the basis of the biomarker (Weeks, J. M. 1995; Van Gestel, CAM & Van Brummelen 1996; Ricketts, H et al. 2004). The stress burden on an organism induced by the contaminants emerges as biological response which is the fundamental principle for developing toxicity assays in order to derive threshold levels for chemicals, mainly the concentration at which the chemical is toxic or has no effect. Various factors such as physico-chemical properties, exposure dose, duration (acute, chronic), exposure pathway (dermal contact, ingestion, inhalation), or life cycle parameter that has been exposed (adult, juvenile, gender), influence the threshold value. Recently, numerous acute toxicity assays have been standardised to observe the environmentally induced responses (Pauwels et al. 2013). Currently, developments such as molecular techniques assist ecotoxicologists to study the mode of action of the contaminant at earlier 31

32 stages; such discoveries are real time PCR, pyrosequencing c DNA subtractive libraries, DNA chips, advanced bioinformatics programs, statistical data analysis and gene expression studies (Brulle et al. 2010b). Gene expression studies are more beneficial because help us identify the effects of contaminants in terms of gene expression patterns which are responsible for the phenotypic variations of an organism in later stages (Brulle et al. 2008; Brulle et al. 2010b). The salient feature of fundamental ecotoxicology is the test organism involved in experimental studies because it is often unexposed to a contaminated site. Thus the biological response obtained from such an organism compared with a control, which provides us with a more realistic picture on the impact of contaminants in particular individual rather than a natural genetic variation (Pauwels et al. 2013). The term phenotypic plasticity concerns the power of an individual to sustain stress conditions which is commonly denoted as acclimatisation in ecotoxicology (Pigliucci 2005; Van Kleunen & Fischer 2005; Ghalambor et al. 2007). This is the root cause for the emergence of the term biomarker. Various biomarkers in earthworms are tabulated below (Table 2.3). 32

33 Organism Eisenia fetida Andrei (Adult worms) E.fetida Andrei (Adult worms) Contaminants Concentration and duration of exposure Carbaryl 12, 25 and 50 mg kg -1 2, 7, 14 days Benzo(a)pyren e 0.05, 1, 100, and 1000mg kg -1 2, 7, 14 and 28 days E.foetida Chlorpyrifos 2.96±0.39 and 2.33±0.39 mg/ml 12,24,36 and 48 hr Soil condition Biomarkers Effects Reference Artificial soil Soil moisture = 35% Temp=20±1 degree ph= 6.5±0.5 continuous light Artificial soil Soil moisture = 35% Temp= 20±1 c ph= 6.5±0.5 Continuous light Filter paper contact test (OECD, Method 209) 33 AchE, MROD, NADH Red, NADPH Red, Catalase, GR, GST, LP and LPI, Total GSH, % GSSG AchE, MROD, NADH Red, NADPH Red, Catalase, GR, GST, LP and LPI, Total GSH, % GSSG AchE Significant inhibition in the enzyme activity mainly MROD, NADH Red and NADPH Red) but no dose dependent response. All the treatments inhibited AchE activity in response to time. Reactive oxygen species production is induced by B(a)P exposure, MROD activity was initially initiated, later decreased at higher concentration, lipid peroxidation was observed, catalase production was initiated but no dose dependent response. Significant effect on the AchE activity (>60%) Abnormal morphology such as necrosis, structural damage, bloody lesions, rectal area detachment. (Ribera et al. 2001) (Saint- Denis et al. 1999) (Reinecke, S & Reinecke 2007a)

34 Drawida willsi (Juveniles) Lumbricus rubellus (Adults) Aporrdctod ea caliginosa (Adult and Juvenile) Butachlor, Malathion and Carbofuran 1.1and 2.2 mg kg -1 (Butachlor and carbofuran) 2.2 and 4.4 mg kg -1 (Malathion) 1 to 105 days Pyrene 10, 40, 160, 640, 2560mg kg -1 Diazinon:Basu din 600W, 60% a.i Chlorpyrifos:L orsban 40 EC, 40% a.i Diazinon=12 and 60mg kg-1 Chlopyrifos=4 28mg kg days Natural soil Soil Moisture= 20% Temp= 25±2 C ph= 6.8 Artificial sterilized loam soil Soil Moistur= 80% Temp= 15±1.5 C 16:8 hrs, L:Dcycle Natural soil Soil Moisture= 25% ph= Temp=20C Continuous light AchE EROD Catalase NRR assay, ChE, GST Butachlor showed no effect on AchE activity. 41 to 47% inhibition in AchE was after 9 and 12 days of Malathion exposure, 54 to 62.9 % inhibition in AchE was seen in carbafuron treatment after 12 days. No significant effect on EROD, Catalase activity was inhibited at higher concentrations of 640 and 2560 mg kg -1 with respect to control. Juveniles are vulnerable, diazinon about 12 mg kg-1 inhibited 75% of ChE, whereas 90% inhibition was observed in 60 mg kg % inhibition in ChE activity was observed in chlorpyrifos at 4 to 28 mg kg -1. Lysosome membrane stability was disturbed by both diazinon and chlorpyrifos in adult worms. ChE, NRR assays act as suuitable biomarkers for (Panda & Sahu 2004) (Brown et al. 2004) (Booth, LH & O'Halloran 2001) 34

35 the pesticide. Eisenia fetida (Adult) Eisenia fetida (Adults) chlorpyrifos Lindane, Deltamethrin 1.5 Kg per hectare 6 months Lindane 20, 50, 80, 120, 150 mg kg -1 Deltamethrin- 5, 25, 50, 100, 150 mg kg days and 42 days Microcosm experiment Soil moisture= 57-63% Temp=20C Humidity=60% Continuous light Artificial soil Humidity= 70-90% Temperature= 20±1C Light= Lux Continuous light ChE NRR assay Growth, cellulase activity Higher concentration say 8.0µg/Kg inhibited the growth of the organism. ChE activity was inhibited but no clear dose response. Lysosome membrane stability decreased with increasing dose. Only in acute dose dependent toxic effect was observed in both the compounds (Growth and cellulase). Mortality was higher in lindane. Both acute and sub chronic effects were severe in lindane. (Reinecke, S & Reinecke 2007a) (Shi, Y et al. 2007) Eisenia fetida (Adults) Eisenia fetida (Adults) Benzo(A)Pyre ne and pyrene PFOS, PFOA 10-6 to mg ml hours whole body exposure PFOS- 0,100,160,256,4 10,655,1050mg Filter paper assay (OECD ) Artificial soil Soil moisture=35% ph= 6.5±0.5 Cytochrome P450, antioxidant enzymes Growth and mortality Test concentration was sensitive to P450, SOD, POD, CAT didn t show clear dose response, decrease in SOD activity failed to protect against BaP toxicity. Mortality was higher in PFOS than PFOA LC 50 PFOS=405.3mg (ZHANG, W et al. 2009) (Joung et al. 2010) 35

36 Eisenia Andrei (Juveniles) Eisenia fetida (Adults) Eisenia foetida (Adults) Organophosph ate pesticide Azinphos methyl kg -1 PFOA- 0,500,750,1125, 1690,2530 mg kg days mg kg -1 (Conducted) mg kg -1 (Selected) 2 weeks 12 weeks PFOS 0-100mg kg hrs, 14 days Chlorpyrifos Fenvalerate 10,20,40mg kg -1 1,3,5,7 days Alternate light and dark pattern Artificial soil Soil moisture=35% ph= 6.5±0.5 Alternate light and dark pattern Filter paper assay Artificial soil test Natural soil Temp= 20±1 C Alternate light and dark pattern ChE, Behaviour, NRRT, Reproduction Lethality, Avoidance behaviour Cellulase, SOD and CAT kg-1 PFOA= mg kg-1 Unable to avoid azinphos methyl in soil, ChE activity was inhibited at all concentrations along with enzyme activity. NRR was much lower than that of control. LC 50 Values Filter paper- 13.6mc*cm -2 Artificial soil mg kg-1 Natural soil mg kg-1 Avoidance was seen at 160 mg kg-1 Cellulase activity was inhibited by both the pesticides exposure. SOD was initially inhibited but recovered after 3 days. CAT activity was initially increased but later reduced. (Jordaan, Reinecke & Reinecke 2012) (Xu, D et al. 2011) (Wang, J-h et al. 2012) 36

37 Eisenia fetida Eisenia fetida (Adults) Eisenia foetida (Adults) Cadmium and Pyrene Phenanthrene, Pyrene Environmental concentration Cadmium 2.50 mg kg -1 Pyrene mg kg-1 8 weeks 0.5, 2.5mg kg -1 Phenanthrene 50, 100 mg kg -1 Pyrene 4 weeks PFOS 0,10,20,40,80,12 0 mg kg days Artificial soil Temperature= 20±2 C WHC= 40% 12:12 Alternate light and dark pattern Microcosm Soil moisture= 60-73% Temperature= 20C Humidity= 60% Artificial soil experiment Soil temp= 20±1C 12:12 alternate ligh and dark pattern Artificial soil Temperature= 20±2 C WHC= 40% 12:12 Alternate light and dark pattern Cytochrome P450 GSH, SOD, CAT Growth, MDA, antioxidant enzyme, NRRT, sperm count, Gene expression Antioxidant enzymes, DNA damage Cadmium copresence increases the toxic effect of pyrene. Initially antagonistic on CYP450, GST, SOD later recovered. CAT activity was increased in later duration. Phenanthrene and pyrene increased the transcript abundance of annetocin and TCTP genes, clear dose response in NEET, growth and antioxidant enzymes activity was much reduced. Sperm count was also decreased in test concentrations PFOS induced oxidative stress in earthworms. DNA damage was visible in higher concentrations. Enzymes like SOD, Peroxidase, CAT, Glutathione peroxidase were inhibited at later duration. (Yang, X et al. 2012) (Wu et al. 2012) (Xu, Dongmei et al. 2013b) 37

38 Apporectod ea caliginosa Eisenia fetida (Adults) PFOS, PFOA Benzopyrene, Pyrene 1,100,500 mg kg days 1.0mg kg -1, 4 weeks Haplic luvisol 2% sand, 81% silt, 17% clay ph= 6.8 OC= 11.2 mg/g 1.93% organic matter WHC=40% Pollution experiment recommended by OECD with PAH contaminated soil. Survival Gene expression Decrease in survival rate and weight reduction initiated above 100 mg kg -1 concentration. PRDX was up regulated in earthworm exposed to treatment, which acted as a valuable biomarker. (Zareitalaba d et al. 2013a) (Lin et al. 2013) Table 2.3 Important biomarkers in earthworms exposed to various persistant organic contaminants 38

39 2.4 Biomarkers in earthworms Role of antioxidant enzymes in defence mechanism Oxidative stress is the condition in the cell, induced after exposure to organic and inorganic pollutants. This condition generates free radicals such as hydrogen peroxide (H 2 O 2 ), superoxide (O - 2 ) and hydroxyl (OH ) mainly as metabolic by-products of xenobiotics in cellular defence system as in the aquatic bryophyte Fontinalis antipyretica Hedw (Dazy, Masfaraud & Férard 2009). The balance between prooxidant endogenous and exogenous factors (i.e. environmental pollutants) and antioxidant defences (enzymatic and nonenzymatic) in biological systems can be used to evaluate toxic impact under harsh environmental conditions, especially oxidative damage caused by different classes of chemical pollutants. In order to tackle the excessive generation of ROS (reactive oxygen species), cells themselves release antioxidant enzymes like superoxide dismutase, catalase, glutathione peroxidase and glutathione reductase along with small molecules (for instance glutathione) to avoid oxidative damage to lipids, proteins and DNA (Valavanidis, Fiotakis & Vlachogianni 2011). The response to oxidative stress activates superoxide dismutase to metabolise the superoxide anion (O 2- ) into molecular oxygen and H 2 O 2. Later the effectiveness of these products is destroyed by catalase and this avoids the incidence of fatal injury. The glutathione reductase enzyme performs a critical role in cellular defence by reducing glutathione in the oxidised form (GSSG) to GSH (reduced and active form). A study on exposure to toluene, ethylbenzene and xylene in earthworms (Eisenia fetida) caused a stress response of the superoxide dismutase (SOD), guaiacol peroxidase (POD), catalase (CAT). It was initiated during the early stages and declined when dosage increased (Liu, Yao et al. 2010). A study by (Wu et al. 2012) reported that superoxide dismutase was induced during the early period on the 2 nd and 7 th days of phenanthrene exposure in Eisenia fetida while its activity decreased with prolonged exposure, and similarly deactivated catalase. Antioxidant activities of the enzymes - superoxide dismutase, peroxidase, catalase and glutathione peroxidase - were initially activated and later inhibited in earthworms exposed to PFOS in artificial soil. This provided evidence for oxidative DNA damage being done to them (Xu, Dongmei et al. 2013c). Oxidative stress can be expressed in many forms, for example lipid peroxidation, DNA damage, ion loss, and failure in protein synthesis. All these symptoms lead to 39

40 complications like cancer in higher organisms (Collins, Andrew & Harrington 2002; Mittler 2002). Injury in base pair, loss of nucleotide, and alteration in nucleotide are dangerous outcomes that are accelerated by ROS (Cooke et al. 2003). Apoptosis is initially observed in the earthworms but continuous exposure to pollutants ends in necrosis (Song et al. 2009). A complex mixture of insecticides and fungicides produces compound effects causing neurotoxicity, also inducing oxidative stress in the earthworms (Schreck et al. 2008). The enzyme that exists in cytoplasmic matrix is GST which facilitates the binding of reduced glutathione with the reactive by-products expelled at the time of DNA dissipation due to oxidation (Leaver & George 1998). The activity of GST is raised or suppressed depending on the toxicant treatment. Experimental results have indicated that GST activity was greatly arrested by PAH (Otitoloju & Olagoke 2010). Interestingly, Srivastava et al. (1999) illustrated that conjugated metabolites transform into a powerful attacking force that fights GST and obstructs the normal function of GST, especially B(a)P metabolites that can restrict the function of GST isoenzymes Detoxification of organics by biotransformation enzymes Survival of earthworms under the burden of carcinogenic and narcotic PAHs was noted due to their capacity to modify these compounds and excrete them from their bodies (Milligan et al. 1986). Detoxification of organics comprises two phases in eukaryotes. The phase I detoxification process links with the cytochrome P450 enzyme system and helps in adding a functional group, such as hydroxyl or sulphonyl, to non-polar compounds. The metabolites of phase I reactions are more fatal than the parent compound. Phase II detoxification enzymes, such as glutathione S-transferase (GST), facilitate binding of large polar, water-soluble moiety to the products of phase I metabolism, and advance the removal of compounds (Donald debethizy & Hayes 1989; Brown et al. 2004). Various studies suggested that xenobiotics in many species were metabolised by cytochrome P450 enzymes (1, 2, 3 and 4) (Roos, PH et al. 1996; Solé et al. 1996; Lee, R. F. 1998; Van der Oost, Beyer & Vermeulen 2003). The cytochrome P450 system of the oligochaetes Eisenia fetida (tiger worm) and Enchytraeus crypticus (pot worm) was estimated using ethoxy-, pentoxy- and benzoxyresorufin as substrates for monooxygenase activity in the whole body microsomes of these earthworms. This study suggested that exposure of the earthworms for up to four weeks to 100 mg fluoranthene or benzo[a]pyrene kg -1 soil (dry weight) did not induce significant alterations in the activity of these monooxygenases. However, short-term exposure to B(a)P by feeding 40

41 retarded the EROD activity significantly by 45%, but did not influence PentROD activity. After long-term (8 weeks) exposure to B(a)P in the agar-agar medium EROD activity was not altered but PentROD had declined to zero. In both species cytochrome P450 and NADPHcytochrome C reductase activities were reported. Since benzo[a]pyrene failed to induce EROD activity in the earthworm species Aporrectodea caliginosa and Lumbricus rubellus, it was proposed that it is not practical to use this as a biomarker for B(a)P exposure in earthworms (Booth, LynnH, Heppelthwaite & O Halloran 2002). In contrast, phase I enzyme activities and lipid peroxidation rates were proportional to both dose of B(a)P and duration of exposure in Eisenia fetida. These results suggested that B(a)P could be metabolised either by P450-dependent activities or by the generation of free radicals. The activities were highly active as significant modifications were detected at the lowest concentration of B(a)P (50 μg kg 1 soil) (Saint-Denis et al. 2001). Detailed reviews reported the role of three mammalian glutathione transferase (GST) families, as detoxifiers of electrophilic xenobiotics and they also made inactive the endogenous secondary metabolites formed during oxidative stress (Hayes, Flanagan & Jowsey 2005). Investigation of GST activity on earthworms when exposed to xenobiotics indicated that GST was non-inducible in most cases (Stokke & Stenersen 1993; Saint-Denis et al. 2001). In the earthworm, however, most of the total GST activity was estimated in the gut, which is represented by chloragogen cells, a part of the coelomic epithelium in the intestine (Saint-Denis et al. 1998). Advanced transcriptomal studies successfully identified the genes responsible for the detoxification of xenobiotics from different chemical classes: inorganic (the metal cadmium), organic (the polycyclic aromatic hydrocarbon fluoranthene), and agrochemical (the herbicide atrazine) in Lumbricus rubellus (Owen et al. 2008). One analysis that integrated sub-proteomics, bioinformatics and biochemical assays reported that L. rubellus GST is a successful biomarker. It possesses a range of GSTs related to known classes, indicating evidence of tissue-specific synthesis (LaCourse et al. 2009) Biomarkers at cellular level In the toxicological perspective the coelomic fluid of earthworms is an interesting area for developing new biomarkers for exposure and effect. Once an organism comes in contact with the toxicant the cells (coelomocytes) play a vital role in defending against foreign bodies by igniting the defence system of the exposed organism (Cooper, Edwin L, Kauschke & Cossarizza 2002; Engelmann et al. 2002; Reinhart & Dollahon 2003). The two main components of coelomocytes are: firstly, amoebocytes that arise from the mesenchymal 41

42 lining of the coelom; and secondly, eleocytes (chloragocytes) which cast off from the chloragogen tissues that enclose blood vessels and intestine in coelomic fluid (Affar et al. 1998; Hamed, Kauschke & Cooper 2002). Coelomocytes are very essential for animals since they protect organisms from fatal consequences. In Eisenia fetida nearly 5 types of cells have been reported, and these are leukocytes type I (basophilic) and II (acidophilic), granulocytes, neutrophils, and eleocytes (Valembois et al. 1985; Calisi, Antonio, Lionetto & Schettino 2009). Biomarkers Types Analytical techniques Eleocyte riboflavin concentration General biomarker of exposure Flow cytometry species Dendrodrilus rubidus References (Plytycz et al., 2007) Lysosomal membrane stability Granulocyte Morphometric alteration Gene expression General biomarker of exposure and effect General biomarker of exposure and effect Specific biomarkers of exposure Neutral red retention assay Diff Quick stain Real-Time PCR Lumbricus spp. Eisenia spp. Aporrectodea caliginosa Eisenia fetida Lumbricus terrestris Eisenia fetida (Plytycz et al. 2007) (Calisi, Antonio, Lionetto & Schettino 2009; Calisi, A., Lionetto & Schettino 2011) (Brulle et al. 2010b) Table 2.4 Biomarkers at cellular level in earthworm coelomocytes (Lionetto, MG, Calisi & Schettino 2012). The immune component of the earthworm cells exhibit auto fluorescence as recorded in eleocytes of Allolobophora chlorotica, Dendrodrilus rubidus, Eisenia fetida, and Octolasion sp. (O. cyaneum, O. tyrtaeum tyrtaeum and O. tyrtaeum lacteum) but not in Aporrectodea sp. (A. caliginosa and A. longa) and Lumbricus sp. (L. castaneus, L. festivus, L. rubellus, and L. terrestris). The reason for this fluorescence is due to riboflavin storage in the cell as identified in Eisenia fetida coelomocytes (Cholewa et al. 2006; Plytycz et al. 2007). Thus the immunity in invertebrates correlates with vertebrates facilitated by vitamin B 12 riboflavin (Verdrengh & Tarkowski 2005). This could act as a candid biomarker of exposure 42

43 to environmental contaminants like POPs because their applicability for metal contamination is widely acknowledged. Lysosome membrane stability represents the biological response from an organism exposed to environment stress (Svendsen, Claus et al. 1996; Maboeta, Reinecke & Reinecke 2003; Svendsen, C et al. 2004). Once an organism comes into contact with the contaminants the first reaction occurs in the lysosome which further leads to pathological consequences for earthworms (Moore, Icarus Allen & McVeigh 2006). A well-established method to calculate the lysosome membrane stability in earthworms coelomocytes is to employ neutral red retention assay (NRRT) mainly in vivo (Weeks, Jason M & Svendsen 1996). Basically, quantification is carried out by observing the time required for 50% of cells to leak the neutral red dye from the damaged lysosome into the cytosol. Sensitivity of lysosome membrane stability as a biomarker of exposure and effect has been acknowledged in various studies that compared other lifecycle parameters in earthworms (Sanchez-Hernandez, J 2006). Semi-field application of lysosome membrane stability as a cellular biomarker was well examined using Aporrectodea caliginosa after exposure to organophosphate pesticides (Lynn H. Booth 2001). Apart from earthworms, lysosome membrane stability has been investigated in ragworm (Hediste diversicolor), and here B(a)P significantly affected the integrity of lysosome. This windicated that coelomocytes were highly vulnerable to PAHs attack (Catalano et al. 2012). More investigations are being carried out on searching for pollutant-induced morphometric alterations in Eisenia foetida granulocytes. The aim is to apply them in soil monitoring. Reports indicated that methiocarb significantly induced the enlargement of granulocytes in earthworms (Calisi, Antonio, Lionetto & Schettino 2009; Calisi, A., Lionetto & Schettino 2011). The link between cellular biomarker response and biological endpoint was clearly indicated in particular with pollutant-induced granulocyte enlargement in Lumbricus terrestris. It indicated high sensitivity to contaminant exposure, thus proposing possible applications as a sensitive, simple, and quick general biomarker for monitoring and assessment. Diff-Quick stained cells image analysis helped to identify granulocyte morphometric variations in the cells (Calisi, Antonio, Lionetto & Schettino 2009). These large cells functions were related to phagocytosis (Engelmann et al. 2002; Cooper, Edwin L & Roch 2003). The general underlying process in enlargement of cells is malfunctioning in the intracellular osmolarity due to metabolic changes such as improper protein catabolism and ion exchange between the cells. These have been reported in studies using heavy metals and pesticides (Lionetto, M et al. 1998; Scott-Fordsmand & Weeks 2000; 43

44 Sanchez-Hernandez, JC 2006). It has been speculated that pollutants (cooper sulphate or methiocarb) also affect the cytoskeleton by reducing the microtubule formation, indicating that xenobiotics target the cytoskeleton which us vital in regulating the cell volume. These outcomes are supplementary to morphometric alterations observed in earthworms granulocyte after pollutant exposure (Gomez-Mendikute & Cajaraville 2003). Granulocytes play a vital role in the immune system of earthworms, so any malfunctioning in their activity leads to fatal consequences (Cooper, Edwin L, Kauschke & Cossarizza 2002). One should also note that the immune system is highly susceptible to attack by pollutants and endangering an organism s health. Even a delicate alteration in any of the immune components serves as a prognostic tool to identify the ill effects induced by contaminants (Lionetto, MG, Calisi & Schettino 2012). One study conducted on Lumbricus terrestris indicated that the enlargement of granulocyte was consistent with the impact on lifecycle parameters such as reproduction (Calisi, A., Lionetto & Schettino 2011). This entails estimating contaminant-induced stress at the cellular level in a comfortable, responsive, and cost effective way. Moreover, it renders a sensitive generalised response to pollutants that can pool the combined impact of multiple contaminants available in the soil. Some recent advances include exploitation of earthworm coelomocytes for transcriptomic studies to identify the up regulated and down regulated genes exposed to metal contamination (Brulle et al. 2008). An extensive review examined how transcriptomics improves our understanding of the molecular-genetic responses of three contrasting terrestrial macro invertebrate taxa (nematodes, earthworms, and springtails) to inorganics, organics, and agrochemicals (Brulle et al. 2010b) Genotoxic biomarkers An ecotoxicological approach on coelomocytes helps to determine the genotoxic effect of both inorganic and organic contaminants in earthworms. The structure and integrity of DNA is highly susceptible to POPs and metal exposure. It is critical to understand the pathway of DNA damage because they encode the genetic instructions used in the development and functioning of all known living organisms. They are also employed to estimate earthworms health. Limitation exists in the studies that assessed the genotoxicity of contaminated soil on earthworms (Button et al. 2010; Espinosa-Reyes et al. 2010; Klobučar et al. 2011). The main methods used to identify the genotoxicity of chemicals in the environment were single cell gel electrophoresis (or comet assay) and micronucleus tests, because they are fast, sensitive and more reliable than other assays. Comet assay estimates 44

45 ruptured DNA in single cells, as single- and double-strand breaks, alkali-labile sites, oxidative DNA base disturbance (Cotelle & Ferard 1999). The procedure in comet assay includes embedding cells in agarose gel on microscope slides and lysing them with detergent and large amounts of salt. Slides are then immersed in an alkaline solution to induce breakage of DNA at alkali labile sites. Under electrophoresis in alkaline conditions, disturbed DNA shows elevated movement from the nucleus to the anode. The ruptured DNA moves further in the electric field, and the cell then resembles a comet with a brightly fluorescent head and tail. The degree of migration is related to DNA damage and efficiency is high since the assay is able to detect even the smallest DNA damage (Tice et al. 2000; Lee, Richard F. & Steinert 2003). Few representative studies have been carried out in earthworms to determine the DNA damage in coelomocytes. On the other hand micronucleus tests are effective in measuring the chromosomal loss and breakage during cell division in vertebrates and invertebrates. Micronucleus are non-integrated chromosomal breakages in daughter nuclei, which was first identified in earthworms exposed to B(a)P by (Sforzini et al. 2012a). DNA damage was observed in isolated coelomocytes treated with hydrogen peroxide and cadmium, and also in coelomocytes from earthworms exposed for up to 21 days to soil containing polycyclic aromatic hydrocarbons (Di Marzio et al. 2005). Genotoxicity was investigated in earthworms exposed to polluted soil from a coking plant full of PAHs after 4 and 10 days. The results indicated that PAHs are genotoxic to earthworms even at acute exposure (Bonnard et al. 2009). Interestingly, even the earthworms body parts exhibited different susceptibilities to DNA damage upon exposure to B(a)P and lindane in Apporrectodea longa. Earthworm intestine was significantly (P<0.0001) more susceptible than crop/gizzard to B[a]P and/or lindane (Martin, FL et al. 2005). The behaviour of PFOS is similar to PAHs in inducing DNA damage reported in earthworms from contaminated artificial soil and filter paper assay (Xu, Dongmei et al. 2013b). The field application of comet assay was investigated by applying comet assay to earthworms from a polluted site and laboratory earthworms in Mexico. Here the soil dominant in hexacholorobenzene, lindane and PCBs caused significant DNA damage (Espinosa-Reyes et al. 2010). The reliability of comet assay was further proved by the test conducted on earthworms exposed to endosulfan, where DNA damage was confirmed in both earthworms and clove plants (Liu, W et al. 2009). 45

46 2.4.5 Biomarkers on blood chemistry In vertebrates, haematology has been much explored by using them to signal environmental stress conditions induced by anthropogenic compounds. But they have not been widely used for invertebrates (Bowerman et al. 2000; Dauwe, Janssens & Eens 2006; Rogival et al. 2007). The role of haemoglobin as an oxygen carrier in earthworms was investigated as early during 1942 using Lumbricus herculeus Savigny (Johnson, M 1942). More recently, Calisi et al. (2011) experimented with heavy metals (cadmium, copper, mercury) and reported changes in Hb concentration and its oxidation state in the earthworm Lumbricus terrestris. Employing an integrated approach focusing on heavy metal pollution, it has been suggested that haemoglobin is much sensitive in Lumbricus terrestris (Calisi, A. et al. 2013). MetHb induction could be regularly applied for assessing environmental monitoring in terms of measurable biological response. In fact, it displayed an extensive alteration following heavy metal exposure. More studies are required to measure and identify the responses specific to xenobiotics in order to evaluate the general health of earthworms from polluted sites. MetHb induction could serve as a suitable prognostic tool of exposure/effect in a multi-biomarker approach in earthworms to monitor soils. Extensive studies have focused on heavy metals effects on haemoglobin and assessing the effects of organic contaminants on invertebrates Behavioural response as biomarkers Behavioural test with earthworms (avoidance) was investigated in detail using uncontaminated, artificially contaminated and originally contaminated soils. A study reported that avoidance behaviour is primarily determined by pollutants, and not by soil chemicalphysical properties (Hund-Rinke, Kerstin & Wiechering, Hendrik 2001). Deviation in normal functions like locomotion, feeding, and sexual interest are expected when an organism is subjected to stressful conditions. Some significant variations are noted in earthworms burrowing capacity, movement over the surface, ingestion rate and avoidance behaviour (Slimak 1997; Jager et al. 2003; Christensen & Mather 2004). These outcomes induce secondary changes in biodiversity. Dominant changes in the locomotion are highly visible in earthworms, where the action is mediated by motor system with octopamine (neurotransmitter), which is part of the central nervous system without any tools (Webb & Orchard 1980; Mizutani et al. 2002). An exhaustive review on earthworm biomarkers highlighted the need to develop biomarkers of behavioural and reproductive disruption with direct implications at individual and population level (Rodriuez-Castellanos & Sanchez- 46

47 Hernandez 2007). Evasive behaviour is the primary evaluation procedure, which can be extrapolated to field conditions (Loureiro, Soares & Nogueira 2005). Even though the method is sensitive, under real field conditions, environmental factors tend to alter their sensitivity. The main targets of pollutants are sensory, hormonal, neurological and metabolic systems, and any damage in these vital processes will lead to intense behavioural changes (Sanchez- Hernandez, J 2006). Significant variation in burrowing was observed by (Capowiez et al. 2003) in earthworms exposed to imidacloprid. Behavioural changes are visible when earthworms come into contact with pesticides, and in case of chlopyriphos there was a null effect (O Halloran et al. 1999; Hodge et al. 2000). Behavioural responses were sensitive to pesticide (imidacloprid) exposure, and chemicals significantly reduced the burrowing capacity of Aporrectodea noctura and Allolophora icterica (Capowiez, Yvan & Bérard, Annette 2006). Recently, signal transduction by xenobiotics has garnered much attention in fish exposed to organochlorines, organophosphates, carbamates and heavy metals. Organochlorines specifically combine with the membrane-bound ouabain sensitive Na + -K + -ATPase, affecting neural transmission while organophosphates and carbamates bind specifically to the membrane-bound enzyme acetylcholinesterase. This also impacts on neural transmission. Since the nervous system is similar to the physiological systems of animals, any malfunction of the nervous system leads to secondary fatal consequences (Bhattacharya 2000). Such identification may be applicable to invertebrates. Biomarkers that estimate impacts on suborganism levels can reveal the relationship between a chemical and its toxic output. These biomarkers are more attractive than the conventional method of contamination assessments, such as mortality and abundance (Booth, LH, Heppelthwaite & McGlinchy 2000) Impairment in reproduction Recently, the most promising earthworm biomarker is the identification of annetocin in E.fetida exposed to varied xenobiotics that are associated with oviparity, cocoon production and weigh loss (Oumi et al. 1994; Ricketts, H et al. 2004; Zheng, S et al. 2008). Most studies which explored heavy metals reported that earthworms treated with lead and zinc showed a 20-fold decline in expression of annetocin gene. Recent data on environmental reproductive xenobiotics contribute to the evidence concerning carboxylesterase overexpression in the male reproductive tract. The main role of carboxylesterase is to detoxify environmental chemicals such as organophosphate pesticides. It is expected that this enzyme will modify physiological functions to defend the male reproductive system against 47

48 xenobiotics that may create altered sperm and maturation differentiation (Mikhailov & Torrado 1999). Potentially, a toxicant disturbs reproduction ability either directly by attacking the process involved in reproduction or indirectly by destroying the metabolism of nutrients (Booth, LH & O'Halloran 2001). Reproductive damage in earthworms is induced by PAHs mainly by reducing the cocoon production and juvenile emergence after 28-day and 56-day exposure in Eisenia fetida (Bonnard et al. 2009). It can be concluded that acute exposure causes serious weight reduction, whereas prolonged exposure tends to affect the earthworms later generations due to a malfunction in reproduction Molecular Biomarkers Reliable, robust, sensitive and specific biomarker is essential for diagnosing the presence of pollutants in the environment and its intensity and effects on living organism. Gene expression is one response of living organisms as a consequence of stress posed by environmental pollutants. This has been explored using molecular markers to diagnose the presence of environmental pollutants and to assess their toxicological effects (Galay-Burgos et al. 2003; Asensio et al. 2007). Toxicity of a pollutant has an impact at the molecular level, i.e. sub-individual level first and as a consequence causes toxicity to the organism and after that it affects the population (Vasseur & Cossu-Leguille 2006). Therefore, using a molecular biomarker, damaging outcomes can be diagnosed early. However, the properties of molecules that we would like to use as biomarkers must be characterised and their responses must be studied. Eisenia fetida has been used in albendazole induced gene expression studies and the authors have suggested that target genes expression could provide information on early identification of pollution of albendazole residue in soils (Gao et al. 2013). Zheng et al. (2008) exposed Eisenia fetida to four different polycylcic aromatic hydrocarbons: BaP, Phe, Pyr and fluoranthene (Flu). They monitored the expression of annetocin and translationally controlled tumor protein (TCTP) by RT-PCR, and found that BaP is a better inducer of these genes than the other PAHs. Furthermore, these authors concluded that mrna transcription level in earthworms was a more sensitive indicator of PAHs exposure than traditional indices. Expression of annetocin and TCTP in Eisenia fetida has been further explored as biomarkers for soil contaminated by phenanthrene (Phe) and pyrene (Pyr); independently and in combination. Using real-time PCR, it has been shown that Phe and Pyr increased the gene expression levels of annetocin and TCTP synergistically (Wu et al. 2012). It has been suggested that ATP synthase b subunit, lysenin-related protein 2, lombricine kinase, 48

49 glyceraldehyde 3-phosphate dehydrogenase, actinbinding protein, and extracellular globin-4 are potential biomarkers for detecting phenanthrene as low as 2.5 mg kg 1 using Eisenia fetida (Wu, S et al. 2013). Eisenia andrei was exposed to polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) contaminated soil and changes in gene expression of specific molecules were recorded. Calreticulin and Hsp70 have been suggested as potential sensitive biomarkers for PCDD/Fs (Roubalová et al. 2014). This study employed specific candidate genes as markers that were tested. However, more specific and highly selective markers are essential as the same set of genes would be induced by similar compounds. For example, PAHs are known to bind the AhR and HIF class of proteins and induce CYP1A1 and CYP1B1 (Moorthy et al. 2002; Hooven et al. 2005; Karp et al. 2005). Similarly, PFOS and PFOA have been shown to induce gene expression through the peroxisome proliferator-activated receptor alpha (PPARalpha) (Bjork, Butenhoff & Wallace 2011; Ishibashi, Kim & Iwata 2011; Stromqvist et al. 2012). It is therefore necessary to explore genes that are highly specific to a single chemical in order to diagnose the health of the soil properly. To conduct this task, high throughput tools are available. We carried out mrna sequencing and in silco transcriptome assembly approach to identify potential biomarkers. We have identified biomarkers for PFOA, PFOS, BaP and pyrene. 2.5 Conclusion Currently, earthworm biomarkers play a key role in assessing the level of pollution in soils. Despite its varied application in the evaluation of contaminants effects the focus has been mainly on aquatic organisms and less so on soil dwelling organisms. Initially, appropriate representative animals should be selected; later characterisation of biomarkers occurs in the field scale which will facilitate rapid pollution assessment. We can conclude that extensive studies on biomarkers were previously restricted to heavy metals, and demand is rising to apply biomarkers to identify the exposure and effect of many xenobiotics in the soil. Secondly, new knowledge gained on biological responses of earthworms to pollutants will establish more practical and applicable biomarkers that document the stress responses in soil organisms induced by contaminants. Earthworms physiological fluids such as coelomic fluids and blood provide an opportunity to discover novel responsive and reliable biomarkers using cytological, biochemical and trascriptomic parameters. Moreover granulocyte morphometric modification informs us on the effect of pollutants which could be included in integrated biomarker 49

50 strategies. It facilitates us to evaluate not only a single contaminant but also for multiple contaminants in the soil. Ultimately, the field level application of earthworm biomarkers is still in the development stage because most investigations have only been conducted in laboratory conditions. Only a few studies have been published on real contaminated sites. Consequently, scientists are enthusiastic about the field of earthworm biomarkers because these can be applied to field conditions as a prognostic tool for evaluating the level and type of contamination. This offers an enriched area of research in soil pollution monitoring and assessment. 50

51 3 Materials and Methods 3.1 Soils and earthworms Soils were collected from Adelaide hills (alkaline and neutral soil). Soil properties have been tabulated below (Table 3.1). Earthworms of species E. fetida used in this study were purchased from local gardening/handyman/building supplies outlet (Bunnings, South Australia). Later the animals were acclimatised to soil by maintaining them in uncontaminated soil for nearly one month. Prior to beginning the experiment, healthy adult earthworms with a well-developed clitellum and a body weight of approximately 0.3 g were separated and depurated for 24 h in order to empty their gut contents (Caceres et al. 2009). Three different soils including 2 natural soils collected from South Australia and one OECD artificial soil were used in this experiment (Table 3.1). Artificial soil was prepared by mixing a ratio of 69% silica sand, 20% kaolinite clay, 10% peat moss and 1% CaCO3 (on a weight basis) according to the standard method advocated in OECD guideline no. 207 (OECD 1984). The ph of the artificial soil was 7.0 throughout the experiment. 51

52 Soil Soil Type ph EC WHC (%) CEC (Cmol/Kg) Total Carbon (%) Total Nitrogen (%) Sand (%) Silt (%) Clay (%) Neutral Chromosol Alkaline Calcarosol OECD soil Table 3.1 Physico-chemical properties of OECD, alkaline and neutral soils 52

53 3.2 Earthworm exposure All the different soils (OECD, neutral and alkaline soils) were spiked using a stock solution of PFOS prepared in solvent acetone to obtain final soil concentrations of 0, 5, 10, 100, 200, 300, 400, 500 and 600 mg PFOS kg -1 dry soil (In order to minimise the solvent effect initially ¼ of the 500 g soil was spiked with chemical in solvent, solvent was evaporated in fumehood and then mixed with ¾ of the remaining soil). The soil samples were thoroughly mixed to obtain a uniform distribution of PFOS in soil and solvent was evaporated by leaving it overnight in the fume hood. Similar procedure was followed for PFOA and PAHs spiking. Spiked soils (500 g) were placed in glass jars (1 L), after which they were maintained at 65% of water holding capacity by adding appropriate amounts of deionised water. The other reagents used were AR grade. All experiments were conducted with three replicates per exposure concentration. Soils without POPs served as controls. The worms were cleaned with deionised water, softly blotted on absorbent paper to remove any excess water and depurated for 24 h before being introduced into the test soils. Ten worms were carefully introduced into each of the triplicate jars containing POPs treated and untreated test soils. The containers were loosely capped with lids that were punched with holes to allow aeration. All the test containers were incubated at a constant temperature room at 20 C, with a controlled light (500 lux)-dark cycle 8:16 h. After 1, 7, 14 days of exposure, live worms were recorded. After 14 days the live worms were separated from soils, cleaned with deionised water and weights were recorded following 24 h depuration. The LC 50 values (concentration of the pesticide that caused 50% mortality) were determined after 14 days of exposure by plotting the log concentration of fenamiphos against the probit-transformed percentage mortality values (USEPA, 2000). 3.3 Cellulase activity Two earthworms from each container were taken on the 14 th day, placed on filter paper (moistened with deionised water) for overnight to facilitate the removal of gut contents, weighed, and homogenised with ice-cold deionised water. After homogenising, the contents were centrifuged for 10 min at 2500 rpm and the supernatants were again centrifuged for 5 min at 3000 rpm. The homogenates were stored at 4 C (Shi, Y et al. 2007). Cellulase activity in the gut was estimated using carboxymethylcellulose (CMC) assay (Ghose 1987). Briefly the assay was conducted with 1ml of CMC-NACL and 0.5 ml of centrifuged homogenate incubated at 50 C for 0.5 h. DNS (3,5-dinitrosalicylic acid) reagent was used to estimate the 53

54 reducing sugar concentration in terms of glucose from the standard curve developed using Bio-Tek synergy TM HT mutli detection microplate reader (Miller 1959). The unit for cellulase activity was milligram glucose released per milligram of protein per hour. Lowry method was utilised to calculate the protein content in the supernatant, using bovine serum albumin as a standard (Classics Lowry et al. 1951). 3.4 Collection of coelomocytes The extrusion method was implemented to collect the coelomocytes from the earthworms. According to the Eyambe et al. (1991) method of extrusion, individual earthworms were rinsed with saline water, and left on the filter paper in order to facilitate the removal of gut content. Later on earthworms were placed in a Petri-dish containing 3 ml of extrusion buffer. The buffer is a mixture of 5.0% of ethanol in 95% saline (0.85g NaCl in 100ml of water) supplemented with 2.5 mg/ml of EDTA, and finally the content was adjusted to ph 7.3 with 1N NaOH. Incubation period lasted for approximately 3 min following which the earthworms were removed and returned back to the soil. The collected coelomocytes were washed repeatedly with PBS at 3000 rpm. The supernatants were discarded and finally resuspended in PBS for further assay (Eyambe et al. 1991). 3.5 Cell viability test Viability of coelomocytes were carried out by Trypan blue staining which differentiates live and dead cell (Eyambe et al. 1991). The coelomocytes were subjected to viability test with 0.4 % of Trypan blue (Sigma chemicals co., Australia). Twenty microliters of coelomic fluid were placed on a hemocytometer (with 1*10 6 cells/ml), and 1 ml of trypan blue solution (0.4%) was subsequently applied. In this protocol, viable cells contain clear cytoplasm whereas the dead cells cytoplasm appears blue. Viability is represented as percentage of live cells based on counting. Staining with trypan blue facilitate us to differentiate between live and dead cell (Eyambe et al. 1991). 3.6 Comet assay Earthworms coelomocytes were harvested using a slight modification of the protocol described by Singh et al. (1988). Briefly, earthworms were put into a 1.5 ml Eppendorf tube and then dipped in an extrusion medium containing 5% ethanol, 95% saline, 2.5 mg ml -1 EDTA, and 10 mg ml -1, and glycerol ether to extrude coelomocytes. Extrusion fluid was 54

55 centrifuged and the supernatant was removed. The coelomocytes were isolated in PBS and washed three times using micro-centrifugation for 3 min at 8000 rpm. About 50 µl cell suspensions were suspended in 75 µl of 0.5% low-melting (37 o C) agarose and layered onto a slide pre-coated with normal melting agarose (60 o C) prepared in phosphate-buffered saline (PBS) and then allowed to solidify at 4 o C. The number of cells per sample was Cells were lysed using freshly prepared cold (4 o C) lysing solution whereby cellular proteins were removed and the damaged DNA was liberated. Subsequently, DNA unwinding was done with an alkaline solution (300 mm NaOH and 1 mm EDTA, ph 12.6) for 30 min. Employing ph of 12.6 helped to distinguish the extended comet tails from the heads easily. Electrophoresis of the DNA was run in the same buffer for 30 min at 300 ma and 25V in an ice bath. Here, the electric field allowed the broken ends to migrate towards the anode. To prevent additional DNA damage, procedures were performed without direct light. For the purpose of neutralisation, the slides were washed with 0.4 M Tris buffer, ph 7.5 and rinsed with ultrapure water. Finally, slides were stained using SYBER green fluorescent dye. Images were analysed using a fluorescence microscope equipped with an excitation filter of nm and a 580-nm barrier filter. Images of 50 randomly selected cells from each treatment were prepared for the analysis. The comet tail lengths (nuclear region + tail) and the width of the comet ''head'' were measured with the comet score software using the Komet 5 image analysis system (Kinetic Imaging Ltd., Belfast, UK). Subsequently, tail length, OTM ("Olive" tail moment (Eq. 2) and % tail DNA (Eq. 1) were obtained using the comet score software: % Tail DNA = I t / I c *100 Eq.1 OTM = Tail DNA% x Tail Moment Length.Eq.2 Where, I t is the total comet tail intensity which is the sum of all pixel intensity values in the comet tail (and I c is the total comet intensity which is the sum of all pixel intensity values in the comet. Tail Moment Length is the relative distance from the centre of the head to the centre of the tail. 3.7 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 oxidising species. There are three species of antioxidants: enzyme systems (GSH 55

56 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 was used to standardise 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 Cu 2+ 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 = S a /S v = n mol/µl or mm Trolox equivalent S a is the sample amount read from the standard curve. S v is the undiluted sample volume added to the wells. 3.8 Lipid peroxidation Lipid peroxidation in earthworm tissue was carried out with the help of a lipid peroxidation kit from BioVision (Catalogue #K : 100 assays). In the initial sample preparation phase, clean earthworms tissue samples weighted around 10 mg has been taken and homogenised 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 ºC for 60 min. Following incubation they were read at 532 nm in a microplate reader. Calculation C= [(A/(mg)]*4*D = n mol/mg Where: A: Sample MDA amount from the standard curve (in n mol) mg: Original tissue amount (10 mg) 4: Correction for using 200 µl of the 800 µl reaction mix 56

57 D: Dilution factor 3.9 Neutral red retention Assay Lysosomal membrane stability was assessed using the neutral red retention time assay described by Weeks, Jason M & Svendsen (1996) with slight modifications. A stock solution of neutral red was prepared by dissolving 20 mg of neutral red in 1 ml of dimethyl sulfoxide (DMSO) and the working solution was prepared by diluting 10 ml of the stock solution with 2.5 ml of PBS. Care was taken that the final concentration of the solvent (DMSO) was kept below a cytotoxic effect level. A 40 µl sample of coelomocytes collected into PBS was placed on a slide and cells were allowed to adhere for 2 min in a humidity chamber at 18º C, and equal volume of neutral red working solution was applied and a cover slip was placed. The slides were incubated in a humid chamber and the neutral red detention was determined under the light microscope (40x objective, Olympus BX41), in 5 to10 min intervals the time at which the dye that had been taken up into individual lysosomes had leached out into the cytosol. The time was recorded when the red coloration of the cytosol was evident in 50% of the coelomocytes (Klobučar et al. 2011) MTT assay The MTT assay was carried out according to Carmichael et al. (1987). Slightly modified from (Maleri et al. 2008). 50µl of micro filtered (0.45mm cellulose nitrate filter), colouring solution (10 mg ml1 MTT (3-(4,5dimethyldiazol- 2-yl)-2,5-dipenyl; Sigma- Aldrich, Australia) dissolved in 20 ml PBS) was added to 50µl of the coelomic fluid. The set up was left undisturbed for 4 hrs, after incubation 50ml of MTT extraction buffer (0.1 N HCl in isopropanol (Sigma-Aldrich), 10% Triton-X; ph adjusted with acetic acid to 4.7) and 50 ml of DMSO was added into every well and the samples were allowed to incubate for a further 3 h. After second incubation the samples were read at 570nm absorption using microplate reader. All readings were corrected by the subtraction of the mean of the absorption of the PBS blank. Tests were considered valid only if the right PBS blank did not differ more than 20% from the left blank LDH cytotoxicity LDH cytotoxicity assay was conducted using LDH cytotoxicity kit from Biovision (catalog #K , 500 Assays). The basic principle is to calculate the activity of LDH (Lactase dehydrogenase) in the cell which was released rapidly upon any damage to plasma 57

58 membrane. Earthworm coelomocytes were collected from all the treatments and control. Collected coelomocytes (2.104 cells) were washed in PBS buffers. The coelomocytes were placed in the microplate. Positive control (1µl LDH) was carried simultaneously to ensure the sensitivity of the kit. 100 µl of LDH reaction mix was added to all the wells, left undisturbed for 30 min at room temperature. After incubation absorbance was measured at 450nm. Calculation Cytotoxicity (%) = (Test sample- Low control) / (High control- Low control)*100 Low Control 100 µl of cells High Control - 100µl of cells combined with 10µl of cell lysis solution (Provided in the kit) Calculation of PFOA and PFOS in soil and animal tissue Soil samples were dried, sieved before analysis. After drying around 15g of soil was taken for analysis. Soil was placed in a centrifuge tube 25ml of extraction solvent (Methanol: ethylcaetate + 7% orthophosphoric acid was added to 40 ml glass vials. The vials were subjected to rotation in rotating shaker for 4 hr and left for settling. Supernatant was allowed to pass through anhydrous sodium sulphate in order to remove trace water content into the samples. 1ml of the supernatant was placed in 1.5ml of HPLC vials for analysis by HPLC- MS. Similarly known amount of earthworm tissue was taken and homogenised with the extraction buffer. Pinch of silica gel was added in order to avoid the disturbances from fat content. The vials were subjected to rotation in rotating shaker for 4 hr and left for settling. Supernatant was allowed to pass through sodium sulphate in order to remove trace water content into another vial, since the amount of PFOS might be low in tissue we carried out a pre-concentration process by evaporating the solvent under a stream of nitrogen gas. The target analytes in this method were analysed by mass spectrometry in negative electrospray mode. For soil samples, the extraction method is based on USEPA SW 846. This involves extraction of the solid phase using a non-volatile solvent system, followed by concentration of the extract. Water samples were analysed directly. The components are separated by HPLC. An aliquot is injected onto a C-18 column and eluted with a short gradient comprised of methanol and an aqueous ammonium acetate buffer. The eluent is introduced into the ESI source and the negative ions selected and detected by MS, operating in SIM mode. Quantitation is performed using the Chemstation Software included with the instrument through the extraction of specific ions: 413m/z for PFOA and 499m/z for PFOS. 58

59 3.13 Extraction of PAHs PAHs from frozen worms and soil were extracted thrice using 10 ml of hexane/acetone mix (1:1 by volume) in an Ultrasonicator (at 12 khz sweep bandwidth; Soniclean, Australia) for 15 min and centrifuged for 10 min later supernatant was transferred to clean vials. The extracts were pre-cleaned with anhydrous Na 2 SO 4. The solvent was evaporated under a stream of N 2 gas, exchanged with 1 ml methanol purified by soil phase extraction with 2 g of aluminium oxide (5% deactivated, upper part) and 2 g of silica (5% deactivated, lower part). PAHs were sequentially eluted with 15 ml of hexanedichloromethane 9:1, and 20 ml of hexane-dichloromethane (4:1). The elute were combined and evaporated to 1mL, prior to analysis of PAHs in HPLC (Agilent 1200, Japan) fitted with a UV detector at 254 nm using C 18 column ( mm with 5 µm packing; Agilent, Australia). A methanol gradient from % was used for separation of PAH compounds at a constant flow rate of 1 ml min -1. Individual phenanthrene, pyrene and B(a)P standards were used as surrogates and the recoveries were more than 95%. The LC detection limit was 50 µg L Avoidance test The influence of PFOS on the earthworms avoidance behaviour in three soils treated with PFOS was studied according to the ISO and ABNT NBR ISO guidelines, using three replicates of plastic square chambers with 500 ml capacity (13 cm 13 cm 5 cm high) for each treatment dose (ISO 2008). The chambers were divided in two equal parts and 100 g of each soil moistened and maintained at 60% WHC max during the previous week were placed in each side of the chambers. PFOS treatment was applied to the soil in just one side of the each chambers (C1, C2, C3, C4, C5) and the other side remained PFOS -untreated-control soil (C0), along with double control chambers (C0 C0). The size of the chambers and the proportion of soil: earthworms were almost the same used by (Das Gupta, Chakravorty & Kaviraj 2011) for acute toxicity test in a 2 day study. The treatments were mg kg -1 technical PFOS, and the double-control were treated with the maximum volume of the acetone solution (360 μl) used in the treatment. The systems remained overnight in fume hood for solvent evaporation, the divider was then removed and six adult earthworms (>300 mg, with clitellum) were placed all together in the slit in the middle of the chambers. All the chambers were closed with perforated plastic film to allow air circulation, and maintained at approximately 22 C under continuous light for 48 h. At the end of the test 59

60 period, the counting of the worms was done on each side of the chambers. According to ISO , the avoidance to the different soil treatments was calculated by counting the mean number of earthworms in each concentration and compared with the mean number of worms in the untreated control soil. Before the placement of the earthworms in the chambers and after their removal, (3 approximately) 3 g samples of the treated soils were removed for monitoring the ph in order to verify other causes of the earthworms behaviour (Sousa & Andréa 2011). The amount of earthworms was converted to percentage of avoidance by the following equation: R (%) = [(C T)/N] 100, where R = avoidance; C = number of worms in the control (C0) condition; T = number of worms in each dose in the same soil; N = total number of worms (2, 3) (4) (5 (1) Figure 3.1 Pictorial representation of the earthworm avoidance test procedures: (1) introduction of the movable wall in the centre of the test vessel; (2) introduction of the soils to be tested; (3) movable wall is removed; (4) placement of the earthworm in the centre of the soils; (5) covering the test vessel with a lid (perforated). 60

61 3.15 Locomotion The methods was modified from (Zheng, R & Li 2009). The locomotion in earthworms was measured after 14 d. The worms were removed from each container and washed in distilled water. They were placed on clean wooden experimental table where there was a red line drew by food colour. A wet print paper (150 mm 150 mm) was put on the table in 1 cm from the red line. After crossing the red line, the earthworms would crawl to the print paper. Initially there was a black line to facilitate easy calculation. As they had dipped some food colour on their body, the trail they had crawled would be marked. The trail that a worm had crawled in one minute was overlaid by a thread. Then the thread was stretched and the length was measured by ruler. Figure 3.2 Locomotion pattern exhibited by both treatment and control worms 3.16 Cast production Cast production measurement was carried out with slight modification by previous methods. Visual assessment was done based on the shape and size at earlier stage later we followed sieve method for proper cast detection (Capowiez et al. 2010). Along with treatment and control we maintained separate positive control container i.e. soil without earthworms. In this study we separated out casts using a set of 2 sieves (diameter of 15 cm and mesh sizes 3.0, 2.5 mm) since earthworm activity may alters the soil granulometry which cause higher amount of soil retention in some sieves (casting) and a decrease in others (soil consumption). All soil from each container, including the soil that adhered to the walls of the dishes which was removed with a sharp knife, sieving was carried out with utmost care to avoid any 61

62 damage to cast. The set of sieves was manually shaken for 20 s optimised on continuous test before caring with the treatments. The soil retained in each sieve was weighed separately. The impact of earthworm bioturbation was then determined by observing the changes in the particle size distribution (PSD), i.e. weight of fresh soil in each sieve minus the corresponding weight of soil for the control soil (without earthworm bioturbation) Wound healing capacity Procedure to calculate was slightly modified from previous study conducted on earthworms (Cikutovic, MA et al. 1999). After exposure, earthworms were collected, cleaned with distilled water, anesthetized worms using 5% ethanol and wounding process done as follows. Under magnifying glassusing small surgical knife, 3 sided shallow cut was made in the earthworms integuments on the dorsal side of the organism at the posterior end. As we draw a cut there were liquid expulsion which was cleaned with cotton, to check whether the incision is in right position we folded back the resultant flap. Later the worms were placed in a plastic container provided with wet filter paper. Same procedure was followed for both treatment and control worms. The set up was left undisturbed for 5 days at 18 C. After 5 days wound healing in exposed and control worms were calculated by observing the flaps. Under dissection microscope observation was made and judgment was made by observing the completely closed flaps in the worms. Non- healed were reported even if there is one flap left open Statistics Each treatment was conducted in triplicate. SPSS 17.0 statistical software was used to analyse the experimental data, and the results were expressed in the form of mean ± SD. Significant differences (P < 0.05) between the treatment groups and the control was determined using the Dunett test Gene expression study Chemicals Heptadecafluorooctane suifonic acid potassium salt (PFOS), perflourooctanoic acid (PFOA), B(a)P, pyrene were obtained from Sigma Aldrich (Australia). 62

63 Animal treatment The oligochaete Eisenia fetida (Bouché 1972) were purchased from Bunnings (Parafield, Adelaide, Australia) and were utilised initially to develop the laboratory cultures. Animals were maintained in compost soil and also supplemented with fruits and vegetable waste, at 25 ±1 º C, 60 to 80 humidity and a 16:8 h L: D (Light: dark) cycle. Adult worms with well-developed clitellum weighs between (range: mg wet weight) were used in the chronic toxicity assays Experimental Design Earthworm, Eisenia fetida was introduced into neutral soil (ph, 7.3) for the differential gene expression studies. Soil was collected from Mount lofty, shade dried and sieved under 2 mm sieve. Assay using pure substance, PFOS, PFOA, B(a)P, pyrene were carried at 10 mg kg -1 concentration; controls were maintained simultaneously. Before introduction of earthworms, soils were artificially contaminated by mixing the soil with individual POPs in acetone solution and shaken end-over-end overnight. Initially 50 g of soil was mixed with the required amount of acetone and later they were mixed with the remaining soil which facilitates homogenisation of spiking. The solvent from the soil was evaporated in fume hood. Prior to introducing earthworms, the soils were wetted with water to 50-60% by mass of the total water holding capacity. Initially 15 worms were released in each container. The containers were added with 40 g of fruits and vegetable every week and also supplemented with powdered pulses. The cycle was continued up to 6 months. Treatments and control were maintained under the same condition RNA Isolation and next generation sequencing Total RNA was isolated using QIAGEN Mini Kit (Cat No: 74104) using manufacturer s protocol. The Eisenia fetida, treated as well as control, were homogenized using Ploytron homogenizer in suspension buffer. The RNA was eluted in elusion buffer and stored in -80 º C till processed for sequencing. The paired-end RNA sequencing was carried out at The Ramaciotti Centre for Gene Function Analysis using Illumina HiSeq Sequence analysis The forward and reverse raw RNA sequence reads were joined separately and analysed using Trinity software (Grabherr et al. 2011) installed in bigmem-1024 server, ersa, Adelaide. Forward and Reverse library was used to assemble transcripts and minimum reads to join k-mers was set at 2 and glue was set at 4. The assembled transcripts were 63

64 translated and longest possible ORFs with >130 amino acids were selected and annotated using stand-alone BLAST (NCBI-BLAST ) using UniProt database. NPKF normalized counts were compared and the transcripts that are more than fourfold higher expressed with the significance of were selected. 64

65 4 Acute toxicity of PFOS 4.1 Introduction During the past 50 years, perfluorooctane sulfonate (PFOS) has been widely used throughout the world in firefighting foams and a variety of household products. Consequently, PFOS has become ubiquitous in the environment and recently recognised as an emerging contaminant of public concern due to its extreme persistence, bioaccumulation and toxic (PBT) nature. Thus, the practical utility of this compound has been firmly confirmed in all activities, ranging from household purposes to industrial applications. PFOS has been detected in human breast milk, blood serum, wild animals and in aquatic fishes especially in their livers (Guruge et al. 2005; Kärrman et al. 2007; Lau, Christopher et al. 2007). Recently, PFOS has been included in the Stockholm Convention on Persistent Organic Pollutants (POPs) which prohibits or restricts their manufacture and use due to their PBT nature. Numerous organic contaminants continue to remain in the soil for longer period once discharged into the soil (Das Gupta, Chakravorty & Kaviraj 2011). Soil supports all forms of livings organisms and hence, the quality of soil is highly essential to maintain sustainable ecosystem. Soil contamination results in deleterious effect to all organisms and the resources generated from these organisms (Wild 1993; Ingham, Moldenke & Edwards 2000). The level of exposure to xenobiotics differs from organisms to organism at different trophic levels. The effect of contamination is massive when the organisms at lower level are exposed because these organisms are consumed by others in different trophic levels leading to biomagnification along the food chain. Thus the contaminant passes from one trophic level to other (Vasseur & Cossu-Leguille 2006). Earthworms constitute groups of animals that belong to a very complex soil food web. They are exposed to contaminants that are directly applied to or reach the soil, either because the soil is their main source of food and the contaminants are absorbed by their body surface 65

66 (Yasmin & D'Souza 2010). Due to their ecological relevance, earth worms have been used as bioindicators and, as they are also biosensors of sublethal concentrations and they could serve as a warning sign for the early effects of soil contamination. E. fetida and E.andrei are the commonly used bioindicators of soil contamination. Determination the bioaccumulation factor of the contaminants (BAF) and the bioassays in order to assess the effects of contaminants on earthworm s reproductive parameters have been proved to be sensitive (Sousa & Andréa 2011), but the behavioural changes have been pointed as useful to detect adverse effects generated by their exposition to sublethal doses (Weeks, J & Comber 2005). These behavioural changes due to the presence of contaminants in soils can be detected, for example, by the avoidance test (Yeardley, Gast & Lazorchak 1996; Schaefer 2003). The main advantages of using the avoidance behaviour test to evaluate the ecological risks are: firstly short duration of the test (48 h); and secondly, the fact that is simple set up (Amorim, MJ, Römbke & Soares 2005). However, so far, the avoidance behaviour of the compost worms Eisenia fetida/andrei has not been included as one of the required tests for regulation of pesticides. Nonetheless use of natural soils as substrates for bioaccumulation of chemicals in terrestrial organisms has been adopted recently as guidelines. Most of the enzymatic activities, such as oxidoreductases (e.g. superoxide dismutase, catalase), transferases (glutathione S-transferases), hydrolases (e.g. acetylcholine esterase) in earthworms are regarded as early warning signals for environmental pollution (Saint-Denis et al. 1998; Saint-Denis et al. 1999; Booth, LH, Heppelthwaite & McGlinchy 2000). Cellulase is one of the important digestive enzymes that are responsible in breakdown of organic matter (Lattaud et al. 1997; Zhang, B-G et al. 2000). Previous studies demonstrated that cellulase activity appeared to be sensitive to pesticide exposure in Eisenia fetida, which can be further explored with respect to other POPs.(Luo et al. 1999). Environmental risk assessment accepts biomarker only when there is a clear link between the responses that comes from different biological organisation of the organism, because they act as an early warning signal for the later effects that are going to appear at population level.(kammenga et al. 2000; Scott-Fordsmand & Weeks 2000; Vasseur & Cossu- Leguille 2006; Gravato & Guilhermino 2009). Biological responses can be collected either from laboratory and field scenarios but responses from target population are easier to acquire under controlled condition rather natural ones (Svendsen. et al. 2007). Currently most of the ecotoxicological studies have focuses on finding the link between responses detected at different biological organisation of the body such as sub cellular (acetylcholinesterase, antioxidant enzyme activity) with ecological relevant parameters namely growth, 66

67 reproduction, burrowing, feeding, survival rate (Engenheiro et al. 2005; Moreira-Santos et al. 2005; Gravato & Guilhermino 2009; Howcroft et al. 2009). 4.2 Materials and Methods Refer chapter Results Survival and growth of earthworms The mortality of earthworms exposed to PFOS in different soils is shown in Figure 4.2. Of the three soils tested OECD indicated the highest toxicity followed by alkaline and neutral soils. After the 14 th day, the highest concentrations causing complete mortality in worms due to PFOS exposure were as follows: 600 mg kg -1 in neutral soil, 550 mg kg -1 in alkaline soil and 350 mg kg -1 in OECD soil. The calculated 14-D LC 50 values for these soils ranged between mg PFOS/kg soils (Table 4.1). However, weight loss appeared to be a more sensitive toxicity parameter than mortality (Figure 4.2). Thus exposure to PFOS led to significant weight loss in worms even at 50 mg kg -1 soil. Eisenia fetida is more sensitive to PFOS in OECD soil when compared to alkaline and neutral soils. Analysis of mean weight change over 14 days by one way ANOVA indicated that the effect of Perfluorinated compounds (PFOS) is significant in higher concentrations in all types of soil when compared with the control (p<0.05). Definite dose response relationship was observed in all the concentrations in all types of soil. Post hoc multiple comparisons indicated a significant difference from low concentration, for example 100 mg kg -1 in both alkaline and OECD soils, but the effect of weight change was significant only after 300 mg kg -1 in neutral soil. 67

68 Figure 4.1, 4.2 Survival percentage in earthworms exposed to PFOS (alkaline, neutral and OECD soils) & Weight loss percentage of earthworms treated with different concentration of PFOS in different soil (Neutral, Alkaline and OECD soil) 68

69 Soil Type LC 50 Value Alkaline soil 366.6±3.9 Neutral soil 446.8±3.3 OECD soil 159.9±2.8 Table 4.1 LC 50 value calculated for PFOS in 3 different soils after 14 days of exposure. Figure 4.3 Lysosome membrane stability in earthworms treated with PFOS in 3 different soils (alkaline, neutral and OECD soils) Lysosome membrane stability Lysosome membrane stability of earthworm coelomocytes measured in terms of neutral red retention time is shown in Figure 4.3. All three soils exhibited a dose- related response, evidenced by a decrease in neutral red retention time and an increase in PFOS concentration. Lysosome membrane stability is highly restricted at higher concentrations ( mg kg -1 ) in alkaline and neutral soil, whereas in OECD soil 300 mg kg -1 concentration showed highest reduction in dye retention time (18 min). The effect of PFOS on reducing the lysosome membrane stability was significant to that of control in OECD soil above 100 mg kg -1 of PFOS (p<0.05), whereas in natural soil a significant effect was visible only after exposure of PFOS to 200 mg kg

70 Figure 4.4 Cellulase activity in earthworms exposed to PFOS in 3 different soils Cellulase activity Cellulase activity acts as a candid biomarker which is responsible for digestion. PFOS exposure resulted in reduced cellulase activity in the worms in all the soils (Figure 4.4). There was a gradual decrease in cellulase activity in the completely exposed organism except for a sudden drop at mg kg -1 of PFOS in the OECD soil exposed organism. A smaller concentration of PFOS at 100 mg kg -1 reduced the cellulase activity to nearly 50%, with the effect becoming more significant at concentrations above 100 mg kg-1 in all worms (P<0.05) exposed to alkaline, neutral and OECD soils. Nearly 90% of cellulase activity ceased at the maximum exposure concentration of PFOS, OECD soil 300 mg kg -1, and natural soil at 500 mg kg

71 Figure 4.5 Avoidance pattern exhibited by earthworms treated with PFOS under 3 different soils. The result was expressed in % Avoidance behavior The included concentrations of PFOS in avoidance tests exceeded the LC 50 value (Table 4.1), since there was no mortality within 48 hours. This outcome indicates a clear rejection of worms for all the soils with evident dose-dependent avoidance behaviour exhibited at all concentrations (Figure 4.5). Significant substrate avoidance occurred in the OECD soil (p<0.05) at 100 mg kg -1 and natural soils (p<0.05) at 250 mg kg -1 compared to that of the control. The avoidance behaviour increased in soils in the following order: alkaline soil< neutral soil<oecd soil. Soil Type Concentration Cocoon production (28 days) OECD Juvenile emergence (56 days) 0 21± ± ± ± ± ±2.08 * ±1.52* 6.33±1.52* ±2.50* 3.66±1.53* 300 1±1.20* 0±0* Neutral soil 0 29± ± ± ± ± ±2.08* ±3.20* 11.33±1.15* 71

72 ±3.05* 8±2.64* ±2.08* 4.33±0.5* ±0.57* 0±0* Alkaline soil 0 34± ± ± ±1.52* ±2.52* 18.66±1.53* ±2.65* 13.66±1.52* ±2.08* 8.66±2.30* ±1.53* 6.66±3.70* ±1.23* 0±0* Table 4.2 Cocoon productions and juvenile emergence on 28 days and 56 days of exposure for Eisenia fetida exposed to a control and different concentrations of PFOS under 3 different soil. *- the mean difference is significant at 0.05 levels by Dunnet t test Reproduction All the control soils showed more cocoons and juveniles than treatments. Of the three soils, OECD soil exhibited a severe decline in reproduction rate compared to the two natural soils (Table 4.2). Post hoc comparison indicated that cocoon production was significant in treatments from the control with no cocoon production occurring at higher concentrations in all three soils. No emerging juveniles were noted at higher concentrations of PFOS (i.e. OECD soil 300 mg kg -1, neutral soil 500 mg kg -1, alkaline soil 500 mg kg -1 ). In the case of natural soils, alkaline soil demonstrated a slightly higher reproduction rate compared to neutral soil, but no significant variation was observed. 4.4 Discussion Mortality and Weight loss Of the three soils tested, PFOS showed the highest toxicity to worms in OECD soil. The highest concentration of PFOS that caused complete mortality in earthworms at the end of 14 days exposure corresponded to 600 mg kg -1 in neutral soil, 550 mg kg -1 in alkaline soil and 350 mg kg -1 in OECD soil. Weight loss appeared to be a more sensitive toxicity parameter since it was significant in worms exposed to PFOS even at less than 50 mg kg -1 in all three soils. Analysis of mean weight change over the period of 14 days by one way ANOVA indicated that the effect of PFOS is significant in higher concentrations in all three soils when compared to the respective controls (p<0.05). Definite dose response relationship was observed in all the concentrations in all tested soils. OECD soil exhibited the strongest effect on weight change. Post hoc multiple comparisons indicated a significant difference 72

73 from low concentration, for example 100 mg kg -1 in both alkaline and OECD soils, but the effect of weight change was significant only after 300 mg kg -1 in neutral soil. PFOS exposure induced mortality in the zebra fish larvae even at 1 mg/l at 132 h (Shi, X et al. 2008). However, in our study a significant increase in mortality was observed only above mg kg -1 in OECD soil, whereas in alkaline soil it was around mg kg -1 and for the neutral soil.(joung et al. 2010). However, our study indicated a dose-dependent negative effect in body mass which is highly significant compared with the control. The soils physico-chemical properties are known to influence the survival of earthworm populations (Grant JR 1955). The observed variation in the toxicity of PFOS to worms among the three soils could be due to their differences in properties. OECD soil proved to be highly toxic compared to the other two soils in the current study. Analyses conducted with endogeic earthworm (A. caliginosa) showed that PFOS is highly toxic with 100% mortality occurring at 500 mg kg -1 (Zareitalabad et al. 2013a). Our investigation also resulted in a similar outcome with 100% mortality observed at 500 mg PFOS/kg soil. Most studies on perfluorinated compounds had focused on aquatic organisms while relatively less information exists on terrestrial organisms (Higgins et al. 2007; Li, MH 2009; Jacobson et al. 2010). The observed differences in PFOS toxicity to earthworms documented in various studies could be due to differences in soil composition, duration of exposure and the nature of worms (Joung et al. 2010; Zareitalabad et al. 2013a). An experiment conducted for 21 days by (Sindermann et al. 2002) reported the LC 50 value for earthworms (Eisenia fetida) which was around 373 mg PFOS/kg soil, revealed a significant variation in physiology, behaviour, and weight loss at a higher concentration ( mg kg -1 of soil). Our study which showed effects even at lower concentration below the LC 50 value also reported significant changes in weight; in particular the effects were predominant in the OECD. Compared to (Xu, D et al. 2011) our study highlighted less toxicity in natural soil than artificial soil. PFOS produced negative effects on the growth and survival of aquatic organisms and after exposure to PFOS significant retardation in the growth in L. gibba was identified. This occurred even at a smaller concentration of 31.1 mg/l. Additionally, studies have reported that PFOS disturbs the normal functioning of target organs mainly by inhibiting the thyroid hormones in fish (Power, D et al. 2001) thus altering the growth. This claim is supported by another study conducted on Monina macrocopa freshwater flea where acute and chronic exposure to PFOS produced significant effect on survival (LC 50 was 27.7 mg kg -1 ) and fecundity (9.3 mg/l) (Lee, C, Kim & Choi 2007). PFOS compromised the health status is some of the invertebrates; prenatal exposure of rats with PFOS (1, 5, 10, 15 73

74 and 20 mg kg -1 ) caused serious developmental defects mainly by lowering the thyroxin level in their body. Invariable PFOS affects the entire organism in the ecosystem. Due to lack of sufficient studies it is very difficult to derive definite benchmark values in terrestrial invertebrates for PFOS toxicity (Beach, SA et al. 2006). Earthworms LC 50 value for PFOS was 373 mg kg -1, and subsequently earthworms displayed significant variation in mortality, burrowing and growth. These calculations are too variable to be used in environmental monitoring which means more detailed studies on the effects of perfluorinated compounds are needed. As per the literature, the prevailing concentrations of perfluorinated compounds in Asia were 0.5 ng/g (Naile et al. 2010). National Institute of Environmental Research estimated the level of perfluorinated compounds in major industrial areas and water basin during 2006 and It reported that the occurrence of PFOS is more than PFOA in sediment, surface water and soil. In reference to the available data, current levels of PFOS and PFOA have no adverse effect on terrestrial ecosystem. Most of the data obtained were from acute toxicity studies, so modified impacts were also expected in sub chronic and chronic toxicity assay which may be more reliable for environmental risk assessment than acute toxicity assays. Some of the clinical symptoms exhibited by higher organism like rat, mice and monkey are neonatal death, weight loss, size variations, enlargement of liver, cancer in testicles, alters fatty acid metabolism(sohlenius, Lundgren & DePierre 1992; Seacat, Andrew M et al. 2002; Seacat, Andrew M et al. 2003) Cellulase activity Cellulase activity is mainly responsible for the degradation of organic matter in earthworm s gut. They also break down cellulosic material and plant matters that are ingested by earthworms. As per the previous study, significant increase in cellulase activity was observed in E.fetida compared to Metaphire guillelmi; in fact the activity was 7 fold higher in E. fetida (Zhang, B-G et al. 2000). PFOS significantly reduced the activity of cellulase in the gut and activity was dose dependent. Of the 3 different soils, OECD soil proved to be best able to decrease cellulase activity compared to alkaline and neutral soils. Such variation in the effects could be due to the difference in how PFOS behave in the soils. A previous study by Shi et al (2007), reported that two different pesticides effect on cellulase activity are highly variable, for example deltamethrin reduced activity while lindane promoted it. The possible reason could be the difference in structure and functional group (Shi et al. 2007). Based on this current study s findings, earthworms are unable to adapt to the situation because PFOS is highly persistent and strongly resists environmental degradation. Cellulase 74

75 has been suggested as a bio-indicator of insecticide pollution in earthworms (Luo et al. 1999), because imidacloprid and RH-5849 were found to retard the cellulase activity of E.fetida. Xiao et al (2006) detected negative effect or earthworm on cellulase activity after their exposure to acetachlor (Xiao et al. 2006). We found that there is a strong positive correlation between weight loss and cellulase activity. This indicates that reduction in weight results decreased cellulase activity, further implying that organisms failing to ingest food results in less cellulase activity which is involved in the digestion process. This correlation is highly significant in the three soils that we have tested Neutral red retention assay According to the review conducted by Moore et al (1982) the most predominant cell organelle involving in intracellular digestion is lysosome, apart from being responsible for various physiological functions in the organism (Moore, Pipe & Farrar 1982). Cell toxicity can be determined by various assays; neutral red retention assay is a firmly accepted technique for quantifying the cytotoxicity of various compounds like pharmaceutical agent, surfactants, food additives, pesticides, solvent and other industrial agents (Borenfreund & Puerner 1985; Fotakis & Timbrell 2006). In this current study we can clearly see a dosedependent response in the decline of lysosome membrane stability when the dose increased in all three soil types (alkaline, neutral and OECD soil). There is a strong positive correlation between weight loss and lysosome membrane stability. This indicated the occurrence of lysosomal membrane damage induced at all exposure concentrations of PFOS. It has been proposed that PFCs mainly induce fatty acid oxidation by acting as peroxisome proliferators (Yang, J-H 2010). Increased membrane fluidity leads to membrane damage due to fatty acid oxidation, which was observed in fish leukocytes after exposing to PFOS, which also confirmed that PFOS has a strong tendency to alter cell membrane properties (Hu, et al. 2003). An earlier review made the point that permeabilization of lysosome initiated cell death pathway under distinct stimuli such as excessive production of ROS. This lethal event causes digestion of vital protein and induce capases activity mainly due to abnormal release of lysosomal proteases (Boya & Kroemer 2008). Evidence also suggest that the amphipatic structure of PFCs could modify the cell membrane potential and trigger the changes in the cytosolic ph (Kleszczyński & Składanowski 2009). Thus any infection in the permeability of the cell membrane could have serious implications on the cell defence mechanism against other xenobiotics. Some factors like high sensitivity towards ph, drought, stress makes NRR assay 75

76 non ideal to apply in field based monitoring program, however these factors are considered to be least important (Svendsen, Claus & Weeks 1997). It is therefore evident that PFOS has strong influence on the physiology of the earthworms Avoidance behaviour Various bioassays were used to determine soil quality and utilised for assessing the risk of various contaminants based on the bioavailable fraction. Sublethal parameters like growth, reproduction were less sensitive compared to avoidance behaviour exhibited by the organism once exposed to contaminants. Earthworms when exposed to PFOS exhibited similar avoidance behaviour pattern in most of the concentration of PFOS in all the 3 soils. One recent study on earthworm exposed to PFOS resulted in significant avoidance at 160 mg kg-1 concentration for both artificial and natural soil (Xu et al. 2011). In contrast, our study showed a vast difference in acute toxicity of PFOS in natural and artificial soils. In artificial soil significant avoidance was noted at 150 mg kg -1 of PFOS whereas in alkaline and neutral soils significant avoidance was noted only at 400 mg kg -1 of PFOS. This could be due to the difference in the bioavailable fraction of PFOS between the natural and artificial soils. Avoidance test has been used widely to identify the effect of pesticides in the soil. PFOS induced some behavioural variation in zebra fish, whereby under acute exposure at mg L -1 PFOS the organism swam faster with increasing dosage (Huang et al. 2010). There has been a demonstration to indicate that soil properties of soil influence much on the escape response of earthworms (E. andrei) and springtails (F. candida), they pointed that avoided soils where with low organic matter content and fine texture (Da Luz, Ribeiro & Sousa 2004). A type of avoidance pattern was therefore observed in the aquatic organism. In the present study the earthworms failed to burrow; instead they remained attached to the rim which demonstrated their aversion to harsh conditions Reproduction rate PFOS has particularly strong sublethal effects on the reproduction rate of earthworms compared to lethal effects. Reproduction is an important life event that occurs in earthworm which helps in population explosion (Kammenga et al. 2003), our study showed that PFOS definitely affects the fitness of the earthworms by altering their reproduction rate ultimately results in change to population dynamics. It should be noted that reduction in cocoon production and juvenile emergence was preceded by the effect of PFOS on the survival of earthworms (Fig 4.1), since the higher concentration indicated poorer survival rate. This 76

77 means that there were less adult in the treatments. Similar effects on reproduction were observed in the study conducted with Enchytraeus aldidus exposed to zinc and cadmium (Novais et al. 2011). Even pesticides produce similar responses in E.aldidus, where their impact on survival of the organism has been documented (Novais, Soares & Amorim 2010). Table 4.2 suggest that PFOS inhibit the emergence of juveniles from cocoon, since coccon production and juveniles emergence were not in proportion compared to control. Similar outcomes have been observed in fish, for example chronic exposure of fathead minnow (P.promelas) to PFOS affected the early life stages namely egg, larvae after 47 days (Drottar & Krueger 2000). PFOS at 0.06 mg/l drastically reduced the reproduction process in fish by inhibiting hatchability, survival, growth and time to hatch. Fish are more sensitive to PFOS exposure than any other invertebrates. In the present study, significant negative effects were observed above 100 mg kg -1 in alkaline soil whereas in OECD and neutral soils it was around 200 mg kg -1 (Table 4.2). The present study indicated that the reproduction process in adults was more sensitive compared to other sublethal responses. McKim (1977) stated that most of the early life stages were more sensitive to xenobiotics than adult ones, and chemicals targeting the endocrine system of organisms may affect the normal reproduction ability of adult ones even at the lower concentration where no visible effects on growth were evident (McKim 1977; Ankley et al. 2001). Even in our study significant negative effects on reproduction was initiated at 100 mg kg -1, whereas there were no significant developmental effects at same concentration. The reduction in the juveniles emergence according to the present data agreed with a study conducted in zebra fish, which showed dose dependent reduction in hatching rate at (1, 3, 5 mg/l of PFOS) (Shi et al. 2008). 4.5 Conclusion The methods employed in this study to evaluate biomarkers were suitable for analysing earthworms sub-lethal responses to PFOS under experimental laboratory conditions. Biomarkers showed biological responses to PFOS exposure which can not be provided by chemical estimation alone. The most predictable elucidation on the impacts of PFOS on earthworms can be done when all biomarkers are used in consolidation. Being sensitive to a lower concentration of mainly PFOS biomarkers like cellulase activity and lysosome membrane stability could serve as an early warning signal of soil contamination rather than mortality and weight change. In addition the results gave important information on how individual earthworms biochemically react to PFOS and demonstrated how this population s variation on long-term exposure can be initiated. Consistent dose-response 77

78 relationships were observed in most of the short-term physiological responses of earthworms exposed to PFOS contamination. Overall the toxic potential of PFOS is much enhanced in OECD soil compared to natural soils. 78

79 5 Acute toxicity of PFOA 5.1 Introduction Of the numerous stable organic perfluorinated compounds (PFCs) that exist, PFOA is the most important one used in various industries and has many commercial applications (Renner, Rebecca 2001; Giesy, John P & Kannan, Kurunthachalam 2002). Extensively utilised in the manufacture of non-stick cookware, PFOA occurs inevitably in DuPont s Teflon products. Being a degradation product of many fluorotelomers, PFOA is highly nondegradable in the environment mainly due to the existence of carbon-fluorine bonds (Prevedouros et al. 2006). Recent studies have reported that environmental media such as air, soil and water absorb these compounds, which can accumulate in wild animals and human beings (Fromme et al. 2009; D Hollander et al. 2010; Lindstrom, Strynar & Libelo 2011). The half-life of PFOA is around 3.8 years in human beings once it is accumulated (Olsen, Geary W et al. 2007). Its occurrence varies from high to low intensity depending on the source of origin; recent studies have reported that during , PFOA occurs in rivers ranging from ng/l (Kunacheva et al. 2012). It has been reported that non-targeted organisms are themselves affected by this, specifically if the general population accumulates PFOA to a certain level, for example 2-8 µg/l (Vestergren & Cousins 2009). PFOA constitute a grave threat to people because they have frequent contact with these compounds mainly through ingestion of contaminated food, water and commercial products (Zhang, T et al. 2010; Zhang, T et al. 2011). These compounds also have higher occupational risks along with environmental exposure. One experiment with PFOA production workers revealed that the accumulation of these compounds in the workers ranged from ,300 µg/l (Olsen, Geary W et al. 2000). A few studies have been done on toxicity of PFOS and PFOA in aquatic organisms. Analyses conducted on Daphnia magna, Monia macrocopa, and Oryzias latipes indicated that PFOS was 10 times more toxic than PFOA. Moreover long-term exposure of fish to these chemicals resulted in their offspring 79

80 being badly affected (Kyunghee et al. 2008). Some of the histopathological damage observed in fish was dose dependent. There is not much evidence regarding the toxicity of PFOA on earthworms. It has been documented that PFOS and PFOA both affect Atlantic salmon s reproduction by physiologically damaging the hypothalamic pituitary gonadal axis (Spachmo & Arukwe 2012). The thyroid hormone is responsible for various functions like growth, embryo development, larval formation, metamorphosis, reproduction and behaviour in teleost fish. According to one study, PFOS and PFOA have the capacity to suppress thyroidal hormone activity, thus disturbing the hormone balance in their bodies (Power et al. 2001). Any agitation in the normal thyroidal hormone function disturbs the health of the individual organism. Ultimately this could mean population destruction in the ecosystem. In our study PFOS and PFOA were employed to examine their role in altering the function of reproductive hormones, mainly androgen and estrogen in rainbow trout and fathead minnows (Oakes et al. 2002; Ankley et al. 2005). Less data is available on the acute and chronic toxicity of PFOA on earthworms. In our present study we conducted acute and chronic toxicity tests using earthworms (Eisenia fetida) to gather more information on the lethal and sub-lethal effects of PFOA in a terrestrial ecosystem. 5.2 Materials and Methods: Refer chapter Results Survival and growth The mortality induced by PFOA in the treatments was calculated after 14 days (Figure 5.2). After 14 days the earthworms from the treatment were removed those without movement were considered as dead. The Figure 5.2 clearly showed that PFOA induced dose dependent death in all the treatment with complete mortality at 1200 mg kg -1 in natural soil; whereas in OECD soil there was complete mortality was observed at 1000 mg kg -1. Based on the result the LC 50 values were calculated for PFOA exposed in 3 different soil condition (Table 5.1). 80

81 Figure 5.1, 5.2 Survival percentage in earthworms exposed to PFOA (alkaline, neutral and OECD soils) & Weight loss percentage of earthworms treated with different concentration of PFOS in different soil (Neutral, Alkaline and OECD soil 81

82 Soil Type LC 50 Value Alkaline soil 823.8±5.3 Neutral soil 894.9±5.9 OECD soil 672.1±7.8 Table 5.1 LC 50 Value of PFOS in 3 different soils (Alkaline, Neutral and OECD) In addtion to survival test we also measured the body weight gain in all the treatments (Figure 5.2). This figure clearly shows that the negative effect on the growth was highly dose dependent. However the degree of weight loss is highly significant to that of control group. The pattern of weight loss is stronger in OECD soil compared to natural soils. A 50% reduction in weight (EC 50 ) was observed at early at 500 mg kg-1 in OECD soil, while in natural soil it was observed only at 700 mg kg Lysosome membrane stability Neutral red retention time in earthworms exposed to different soil PFOA concentration is shown in the Figure 5.3. All three soils exhibited a dose- related response, evidenced by a decrease in neutral red retention time with an increase in PFOA concentration. However the retention time in earthworms exposed to maximum concentration displayed shorter retention time at 800 mg kg -1 in OECD soil and 1000 mg kg -1 in natural soils. It is evident that OECD soil demonstrated more lysosome defects through shorter retention time at smaller concentration than natural soils. The effect of PFOA on reducing the lysosome membrane stability was significant to that of control in OECD soil above 300 mg kg-1 of PFOA (p<0.05), whereas in natural soil a significant effect was visible only after exposure of PFOA at 500 mg kg Avoidance behaviour Avoidance test did not reveal any mortality in the treatments used in both natural soil and OECD soils. Earthworms exhibited clear avoidance behaviour at higher concentration above 300 mg kg -1 in OECD soil and above 600 mg kg -1 in natural soils. It is evident from Figure 5.5 that earthworms showed dose dependent response in avoiding the PFOA contamination. As concentration increased the organism failed to stay in the contaminated 82

83 Figure 5.3 Lysosome membrane stability in earthworms treated with PFOA fewer than 3 different soils (alkaline, neutral and OECD soils Cellulase activity Determining the level of cellulase activity provides us valuable information with respect to digestion in an organism. Figure 5.4 illustrates the effect of PFOA on the cellulase activity of earthworms. The negative response in the activity was significant to that of control (p<0.05) above 400 mg kg -1 PFOA in all the soils. Cellulase activity in worms reduced to almost 80 % at the maximum exposure concentration of PFOA, OECD soil at 800 mg kg -1, and natural soils at 1000 mg kg -1. Figure 5.4 Cellulase activity in earthworms exposed to PFOA in 3 different soils (alkaline, neutral and OECD). 83

84 Figure 5.5 Avoidance pattern in earthworms exposed to different concentration of PFOA in 3 different soils (alkaline, neutral and OECD) Avoidance behaviour Avoidance test did not reveal any mortality in the treatments used in both natural soil and OECD soils. Earthworms exhibited clear avoidance behaviour at higher concentration above 300 mg kg -1 in OECD soil and above 600 mg kg -1 in natural soils. It is evident from Figure 5.5 that earthworms showed dose dependent response in avoiding the PFOA contamination. As concentration increased the organism failed to stay in the contaminated soil and moved towards the control compartments. Avoidance is more sensitive in this study because earthworms in OECD soil showed 30% avoidance at lower concentration 300 mg kg - 1, whereas in natural only at 600 mg kg Reproduction data In the soil test, the cocoon numbers and juvenile numbers were more in control compared all the 3 soils. The OECD soil had a significant effect on cocoon production even at lower concentration at 50 mg kg -1 of PFOA where a significant effect on juvenile emergence was observed. Table 5.2 clearly shows that PFOA significantly reduced reproduction rate of earthworms. Post hoc comparison indicated that the cocoon production was significant from the control, with no coon production at higher concentration in all the 3 soils. No juvenile emergence was noted at higher concentrations of PFOA (OECD soil

85 mg kg -1, neutral soil mg kg -1, alkaline soil mg kg -1 ). The order of PFOA toxicity on reproduction was as follows alkaline<neutral<oecd soil. Soil Type Concentration Cocoon production (28 days) OECD Neutral soil Juvenile emergence (56 days) 0 21± ± ± ± ±0.88* 7.67±0.88* ±0.88* 6±1.15* 300 5±0.58* 3±0.58* 400 1±0.58* 0±0.0* 500 0±0.0* 0±0.0* ± ± ±2.08* 10.67±0.98* ±0.58* 7.00±0.58* 500 9±0.58* 4.00±0.33* 700 5±1.25* 1.33±025* 800 2±0.78* 0.67±0.78* ±0.0* 0±0.0* Alkaline soil 0 34± ± ±0.15* 10.00±0.88* ±0.2* 7.00±0.58* ±1.20* 3.67±0.88* ±0.58* 1.33±0.33* ±0.33* 0.67±0.85* ±0.0* 0±0.0* Table 5.2 Cocoon productions and juvenile emergence on 28 days and 56 days exposure for Eisenia fetida exposed to a control and different concentrations of PFOA in 3 different soils. *- the mean difference is significant at 0.05 levels by Dunnet t test. 5.4 Discussion Mortality and Weight change PFOA in OECD soil is more toxic to earthworms, causing complete death at 700 mg kg -1. Even though they managed to survive in a contaminated environment a significant effect 85

86 on change in body mass was noted early, specifically at 400 mg kg -1 in OECD soil. The LC 50 value for PFOA in OECD soil was around mg kg -1, whereas in the natural soil it ranged between mg kg -1. Similarly, a 96 h test conducted on aquatic invertebrates acquired a LC 50 value ranging from 337 to 672 mg/l (Li, 2009). A slight variation was observed with Eisenia fetida exposed under artificial soil treated with PFOA, and resulted in a higher LC 50 value 100 mg kg -1 for a 14-day acute toxicity test. A varied composition for preparing artificial soil may have been responsible for any changes in the values (Joung et al. 2010). PFOA is moderately toxic to earthworms a finding that is consistent with toxicity studies conducted on aquatic organisms. There is not much evidence on the acute toxicity of PFOA on terrestrial invertebrates. We measured weight change in the earthworms after 14 days exposure. PFOA affected the health status of earthworms as observed in our study. A 50% weight loss was observed at 400 mg kg -1 in OECD soil, whereas in natural soil it occurred at more than 700 mg kg -1. Contradictory reports have been published on PFOA having no significant effect on weight change in earthworms (Joung et al. 2010). However, in one example of a higher plant (Brassica chinensis), PFOA caused 50% decline in survival at a lower concentration, i.e mg kg -1. We can therefore conclude that earthworms are moderately sensitive to PFOA contamination (Zhao, H et al. 2011), and emphasise that toxicity varies with respect to soil types/ properties Some aquatic macrophytes (Myriophyllum sibiricum, M. spicatum) are less sensitive to PFOAinduced toxicity (Hanson et al. 2005), but in terrestrial ecosystem PFOA can induce certain toxic effects on earthworms as shown in our analysis. Even in an aquatic ecosystem PFOA is less effective because one acute toxicity study using fish suggested that PFOA has no effect on the health of fish. Nonetheless it was significant in causing the liver to enlarge (Kyunghee et al. 2008). The data from our study suggests that the fitness of earthworms was disturbed at higher concentrations above 400 mg kg -1 in OECD soil, while in natural soil the impact observed for above 600 mg kg -1. PFOA was slightly weaker than PFOS. This is because PFOA is less adsorptive compared to PFOS (Rayne & Forest 2009; Zareitalabad et al. 2013b). Although lethality is less likely to occur in PFOA they still affect the growth of earthworms. Investigating the immuniotoxic potential of PFOA indicated that the ingestion of PFOA through diet significantly reduced the weight of thymus and spleen in male mice. Furthermore it reduced the mice s weight due to fat being depleted from the body. It has been reported that organisms fail to accept food that is highly contaminated with PFOA (Yang, Q, Xie & Depierre 2000; Yang, Qian et al. 2002). Similarly, this current study showed a dose dependent reduction in body weight which might have occurred because 86

87 earthworms had starved until the end of the experiment. Several studies describe the properties of cationic surface, which adsorbs on the surface of the organism very easily, is also responsible for their higher membrane permeability. We can therefore speculate that soil acute toxicity assay brings the organism into contact with contaminants that can be adsorbed into their surface. A further increase in membrane permeability facilitates the movement of these contaminants into the inner cavity where they can react with enzymes associated with growth, survival, and behaviour leading to acute toxic outcomes. Some ionic liquids induce similar effects in lower organisms like Daphnia magna (Bernot et al. 2005; Pretti et al. 2009). Results of this study suggest that PFOA also induces similar effects in earthworms Cellulase activity Digestive enzymes are highly responsible for the decomposition of plant litter and cellulosic materials in earthworms. The occurrence of a number of enzymes, specifically cellulase in the gut of earthworms, indicates their function in the decomposition of plant litter and other cellulosic materials (Dash 2012). Few reports have been published on cellulase activity acting as a biomarker of insecticide contamination (Luo et al. 1999). Imidacloprid, PH-5849 and acetachlor were found to suppress cellulase activity in earthworms. Similarly, Shi et al. (2007) reported that deltamethrin inhibited cellulase activity whereas lindane promoted it, possibly due their differing chemical structures (Shi, Y et al. 2007). Our study clearly shows that sub-lethal exposure of the PFOA retarded cellulase activity, which indicates PFOA endangers earthworms biochemical metabolism. Indeed, exposure to larger PFOA concentrations causes irreversible damage to a series of enzyme processes and physiological behaviour in earthworms and which eventually decreases their cellulase activity. However, there are no data on the effects of perfluorinated compounds on cellulase activity. Our results showed that acute exposure to higher concentrations of PFOA convincingly inhibited any cellulase activity. We speculate that less cellulase activity at higher concentrations of PFOA may be explained by the organisms being starved for longer periods of time, resulting in poor growth. One earlier study indicated that rats injected with PFDA stopped gaining weight, and weighed less than pair-fed controls along with reduced food intake (Borges et al. 1990). An acute toxicity analysis of PFOA and perfluorodecanoic (NDFDA) acids in male Fischer rats revealed that rats dosed with NDFDA lost half their body weight in 16 days and ate virtually no food from day 7 to day 14. An initial similar dose with PFOA resulted in transient decreases in food intakes and body weight being reversed by 87

88 day 7 (Olson & Andersen 1983). Consequently, PFOA contaminated soils were not preferred by earthworms as a survival mechanism, and they starved for a longer period of time Effect on lysosome membrane stability The most widely used method to identify the effect of contaminants in earthworms is neutral red retention assay, which evaluates lysosomal membrane stability (Weeks, Jason M & Svendsen 1996). Currently, scientists are much interested in identifying the link between lysosome membrane stability and ecologically relevant responses such as reproduction, growth, and behaviour in earthworms (Reinecke, S & Reinecke 1999; Scott-Fordsmand & Weeks 2000; Booth, LH, Hodge & O'Halloran 2001; Reinecke, SA, Helling & Reinecke 2002; Maboeta, Reinecke & Reinecke 2004). Sensitivity of this assay was tested by exposing earthworms to pesticide-contaminated soil, and it was found to be effective even at the field application rate of organophosphate pesticides (Reinecke, S & Reinecke 2007b). Our analysis showed a significantly shorter retention time in PFOA treated animals compared to the control. The retention time in control worms was approximately 60 min which is similar to earlier reports that investigated Eisenia fetida (Reinecke, S & Reinecke 1999; Reinecke, SA, Helling & Reinecke 2002). Earlier investigations reported that PFCs can induce a series of adverse effects at different biological levels, including oxidative stress, DNA damage, membrane instability, suppressed filtration rate, and reduced body weight. Specifically, PFOA was stronger in reducing lysosome membrane stability of green mussels (Liu, Changhui, Gin & Chang 2014). The outcome of elevated membrane fluidity with respect to fatty acid oxidation due to direct interaction with PFCs causes membrane damage (Yang, Q, Xie & Depierre 2000; Parolini et al. 2010). Another analysis confirmed that PFCAs are strong depolarisers of plasma membrane (Kleszczyński & Składanowski 2009). Perfluorinated compounds increase membrane permeability by reacting with membrane lipids (Hu, Wy et al. 2003). It should be noted that lysosome membrane damage is not specific to toxic chemicals because other factors can also induce similar damage (Jordaan, Reinecke & Reinecke 2012). Thus the variation in our experiment s soils may be due to the effect of variation in their physicochemical properties Avoidance in worms Stress induced by PFOA concentration on earthworms is demonstrated by their avoidance pattern in 3 different soils. A rapid sub-lethal avoidance behaviour test served as a 88

89 screening tool in earthworm Eisenia andrei and the isopod Porcellionides pruinosus. The latter were exposed to lindane, dimethoate and copper sulphate, while earthworms were exposed to carbendazim, benomyl, dimethoate and copper sulphate (Loureiro, Soares & Nogueira 2005). The study suggested that concentrations of PFOA having a significant effect on avoidance did not indicate much significant effect for other responses such as growth, mortality, cellulase activity and lysosome membrane stability. Studies suggest that toxicants physiological impacts include disruption of sensory, hormonal, neurological, and metabolic systems. The most frequently denoted links with behavioural disruption were cholinesterase (ChE) inhibition, altered brain neurotransmitter levels, sensory deprivation, and impaired gonadal or thyroid hormone levels with respect to aquatic toxicology (Scott & Sloman 2004). Earlier investigations reported that neonatal exposure to certain perfluoroalkyl acids, PFOS and PFOA, can cause long-lasting aberrations in spontaneous behaviour and also stunt learning and memory functions in the adult animal (Viberg, Lee & Eriksson 2013). It was suggested that the avoidance behaviour is primarily determined by pollutants, and not by chemical-physical soil properties (Hund-Rinke, K. & Wiechering, H. 2001) which was very relevant to our study where avoidance was visible in all 3 soil types. Thus behaviour connects physiological function with ecological processes; behavioural indicators of toxicity are ideal for evaluating the impacts of terrestrial xenobiotics on earthworm populations Reproduction effect Earthworms exposed to PFOA under different soil conditions showed significant dose dependent responses to reproduction. While the lethality of PFOA is small, they do induce powerful sub-lethal defects. Our study indicated that reproduction was more susceptibly sensitive to PFOA exposure compared to other physiological parameters. Furthermore PFOA inhibited cocoon production and juvenile emergence in earthworms. Reproductive effects of PFOA have been reported in higher organisms like humans where significant symptoms were fecundity, delayed puberty and accelerated menopause age in females. The long-term implication in rats because of prenatal PFAS exposure was delay in menarche where people identified a positive correlation with delay and PFAS level in the body (Kristensen, S et al. 2013). All these findings confirm that PFOA has potential as a reproductive toxicant. No evidence has been documented on the reproductive effect of PFOA in terrestrial organisms. Studies on rats have failed to show a positive relationship with reproductive defects and PFOA concentration in the body, whereas in humans PFOA damaged the reproductive system (Butenhoff et al. 2004; York et al. 2010). 89

90 An examination of studies on dioxin reported that an organism s reproductive system appeared to be more susceptible to postnatal exposure rather than prenatal exposure (Mocarelli et al. 2011). By contrast in our experiment earthworms suffered due to prenatal exposure to PFOA. A study involving Sprague Dawley rats showed developmental toxicity during 30 mg kg-1 PFOA exposure however this was proved to be statistically insignificant (Butenhoff et al. 2004). In earthworms PFOA emerged as a strong reproductive toxicant that reduced cocoon production and juvenile emergence at 100 mg kg -1. This is consistent with a study conducted on rats where PFOA at 20mg kg -1 hampered the birth of live foetus (Lau, Christopher et al. 2003). No evidence explains the mode of action of PFOA in inhibiting reproduction; it may occur because previous maternal toxicity arrested the production of viable embryos (Bielmeier, Best & Narotsky 2004). PFOS and PFOA have the ability to activate peroxisome proliferator activated receptor alpha (PPARs) in most animals including mice, rats, common carp and zebra fish. When larger amounts of mitochondrial and peroxisomal lipid metabolism are combined with sterol and bile biosynthesis, leads to the activation of PPARα. This process could be responsible for the developmental toxicity due to PFOA exposure (Andersen et al. 2008) because developmental toxicity in mouse induced by PFOA depended on the expression of PPARα (Abbott, Barbara D. et al. 2007). Consequently, nuclear receptor activation is highly sensitive even at lower concentrations of PFOS and PFOA but there is no evidence for onset of disease. They are clearly responsible for a variety of biochemical and morphological dysfunctions such as impaired metabolism of lipids that alter the cell cycle, and targeting an organism s reproduction system (Andersen et al. 2008). Reproductive changes occurred in earthworms after PFOA exposure provided valuable information on PFOA toxicity in relation to terrestrial ecosystem monitoring. 5.5 Conclusion The methods employed in this study to evaluate biomarkers were suitable for evaluating earthworms sub-lethal responses to PFOA under experimental laboratory conditions. Overall the toxicity of PFOA was enhanced in OECD soil compared to natural soils. This study clearly indicates that PFOA is a potential reproductive toxicant to earthworms even though is less lethal than PFOS. 90

91 6 Genotoxicity of PFOS and PFOA 6.1 Introduction The management of organofluorine compounds has become important in the 21 st century, due to their now pervasive influence as environmental contaminants. Although generally their chemical reactivity is not well understood, many fluorinated organics are biologically effective (Key, Howell & Criddle 1997). PFOS and PFOA are the most popular perfluorinated compounds (PFCs) which are also known as perfluorinated tensides (PFT) (Roos, PeterH et al. 2008). Environmental contamination of persistent PFCs are of concern due to their tendency to bioaccumulate in living organisms especially in blood and plasma, which are exposed to various polluted environments (Hölzer et al. 2008). In the past five decades the concentrations of these compounds in the environment have risen due to the their intensive usage in various commercial and industrial products. Their practical utility has been firmly ingrained in all areas because they are highly prized for their physico-chemical properties that have household and industrial applications. Predominantly aquatic animals and higher mammals have been tested to identify the harmful effects of perfluorinated compounds, and it emerged that PFOS exhibited a complex toxicity on various systems in the body, namely the nervous system, reproductive system, immune system, excretory system (Beach, SA et al. 2006). These compounds are carcinogenic, meaning that they can induce tumours in the liver, pancreas, lungs and breast. While phytotoxicity studies on PFOS and PFOA are few in number, this information on plant toxicity helps to identify these compounds the adverse effects of these compounds on the terrestrial ecosystem (Zhao, H et al. 2011). It is important to know the extent and severity of terrestrial ecotoxicity occurring because agricultural soils are highly contaminated with perfluorinated compounds (Renner, Rebecca 2008). The behaviour of these compounds varies from soil to soil so it is critical to identify their behaviour patterns to establish 91

92 appropriate environmental guidelines on soil quality. Ultimately the combination of various bioassays provides us with information on the complex toxicity of perfluorinated compounds. Risk assessment concerning the impact of potential toxicants on the terrestrial ecosystem is gaining importance internationally (Edwards 2002). Earthworms are sentinel organisms in that they act as a biological piston to facilitate air flow in the soil. In other words they are the key to mineralisation and nutrient enrichment in soil. These organisms endure to a great extent the harsh conditions imposed by anthropogenic compounds (Lanno et al. 2004). All of these qualities make them an excellent candidate for eco-toxicological studies (Spurgeon, David J, Weeks & Van Gestel 2003; Römbke, Jänsch & Didden 2005) which in turn helps us to determine the soil quality. The standardised organism for ecotoxicity test are earthworms, mainly the species E. fetida (Bouché 1972) (OECD no: 207; OECD no: 222). A comparatively recent finding in earthworm biomarkers is identifying a dose-effect relationship at varied levels of functional complexity (Caselli et al. 2006). Bioavailable fraction of a toxicant induces certain physiological changes apart from typical high level variations like mortality and abnormality in reproduction. Currently, these changes in physiological processes are the focus of evaluating the toxicity of anthropogenic compounds (Sforzini et al. 2012b). Over the last few decades important developments have been made with biomarkers; they now have more practical applicability in both laboratory and field conditions (Scott-Fordsmand & Weeks 2000; Spurgeon, David J, Weeks & Van Gestel 2003). The biological response from various degrees of biological organisations considered as a suitable biomarker to assess the toxic environment because each response acts as a biological signal indicating the effect of the contamination stress before mortality occurs in an organism. Consequently earthworms are highly valuable to ecotoxicologists because they act as early warning signals for monitoring contaminated sites (Spurgeon, David J, Weeks & Van Gestel 2003). The addition of genotoxic biomarkers in a risk assessment is closely linked to the genotoxic agent having the potential to cause DNA damage which can be passed on from one generation to the next. Major terrestrial contaminants are mutagenic in nature which means that normal cell division is often disturbed and DNA damage induced (Shugart 2000). Environmental monitoring programs employ the comet assay method to evaluate the extent of DNA damage. Comet assays have evolved as the standard method for evaluating DNA strand breakage. They are ideal for bio monitoring due to their simplicity, high sensitivity, rapid result production and cost effective (Collins, AndrewR 2004). Comet assay (single gel electrophoresis) has been utilised more extensively to identify the DNA injury both for 92

93 vertebrate and invertebrate cells (Singh, NP et al. 1988; Cotelle & Ferard 1999; Rahman et al. 2002; Lee, Richard F. & Steinert 2003; Faust et al. 2004; Siu et al. 2004; Frenzilli, Nigro & Lyons 2009). Recently comet assay has been applied in different earthworm species (i.e. Apporrectodea, Eisenisa, and Lumbricus) to measure the level of DNA damage caused by pollutants from the collected coelomocytes (Salagovic et al. 1995; Verschaeve & Gilles 1995; Reinecke, SA & Reinecke, AJ 2004; Martin, FL et al. 2005; Fourie, Reinecke & Reinecke 2007; Manerikar, Apte & Ghole 2008; Liu, W et al. 2009; Lourenço et al. 2011). In this study we tested the genotoxic potential of PFOS and PFOA using earthworm as a model organism. In earthworm, coelomocytes were used to identify the extent of DNA damage. Studies are limited in regard to the acute toxic effect of PFOS on terrestrial organisms (Xu, D et al. 2011). In this study we used the common model organism - Eisenia fetida - to identify the biological response to PFOS and PFOA treatment. The study aims to reveal the toxic effects of PFOS and PFOA, and provide a scientific basis for the comprehensive evaluation of the effect of perfluorinated compounds in soil ecosystems. 6.2 Materials and Methods Refer to chapter Results The results presented in Figures 6.1, 6.2, 6.3 show the percentage of DNA in the comet tail (tail DNA %) in earthworm E. fetida exposed for 14 days to a wide range of concentrations of PFOS and PFOA. There were significant difference between control and exposed groups. 93

94 (a) (b) (c) Figure 6.1 Comet assay image showing DNA damage in earthworms exposed to PFOS and PFOA,(a) Control, (b) Tail initiation due PFOS exposure, (c) Tail initiation due to PFOA exposure 94

95 Figure 6.2 PFOS induced DNA damage, (a) neutral Soil, (b) alkaline Soil, (c) OECD soil. The parameters shown here is % tail DNA and Olive tail movement 95

96 6.3.1 PFOS induced DNA damage The earthworms mortality was observed in terms of their exposure to the different tested concentrations of PFOS, and especially at very high concentrations. Although the software reports several parameters, data for percentage tail DNA and olive tail movement are presented here as a measure of single-strand DNA breaks/alkali-labile sites. The objective was to evaluate DNA damage in earthworm E. fetida exposed to different concentrations of PFOS and PFOA in 3 different soils, since this is considered to be the most reliable parameter (Collins, AndrewR 2004). The results obtained were highly significant with that of control in PFOS invariably in all 3 soils (p<0.01), whereas for the OECD soil it was observed at 10 mg kg -1, and in natural soil it was induced after 50 mg kg -1. Of these three soils the OECD soil showed DNA damage at around 25% at 300 mg kg -1 concentration, whereas in alkali soil and neutral soil the tail percentage was 15% and 20%, respectively (Fig 6.2 a, b, c). Olive tail movement was strong in the neutral soil, at approximately 30% at 550 mg kg -1 (Fig.6. 2 a), whereas in alkaline soil it was around 25% (Fig. 6.2 b). In the OECD soil olive tail movement was about 23% at maximum exposure concentration (300 mg kg -1 ), shown in Fig 6.2 c. 96

97 Figure 6.3 PFOA induced DNA damage, (a) Neutral Soil, (b) Alkaline Soil, (c) OECD soil. The parameters shown here is % tail DNA and Olive tail movement 97

98 6.3.2 PFOA induced DNA damage As the PFOA concentration increased in the soil, the level of DNA damage also increased in all three soils. The OECD soil showed DNA damage at around 40% at 800 mg kg -1 concentration, whereas in the alkali soil and neutral soil the tail percentage was 25% and 30%, respectively (Fig. 6.3 a, b). Tail initiation is strong in higher concentrations especially in the neutral and alkaline soils, at approximately 40%. Olive tail movement is strong in all 3 soils, with 40% at maximum exposure concentration. Fig 6.3 a, b, c clearly shows that there has been a steep increase in olive tail movement percentage in the OECD soil, mainly from mg kg -1 concentration. The other two soils, in comparison, indicated a gradual increase in the value. There were significant differences in all the treatments compared to the control. The percentage tail DNA is highly significant at 250 mg kg -1 treatment. At 100 mg of PFOA kg 1 dry soil, the olive tail movement revealed a significant response (p > 0.05) compared to the control in the OECD soil. The results also demonstrated that DNA damage caused by PFOA were more serious than that of PFOS. It could be concluded that the PFOA is more mutagenic than PFOS despite the latter displaying a high mortality rate. 98

99 Figure 6.4 Total antioxidant capacity in earthworms exposed to PFOS, (a) neutral and alkaline soil, (b) OECD soil and PFOA (c) neutral and alkaline soil (d) OECD soil. 99

100 6.3.3 Total antioxidant capacity In this study the effects of PFOS and PFOA on E. fetida antioxidant responses were examined. The total antioxidant capacity in earthworms exposed to PFOS and PFOA are presented in Fig 6.4, which shows that it was triggered at an initial concentration and peaked around 300 mg kg -1 of PFOS exposed in neutral and alkaline soils and 200 mg kg -1 in OECD soil. Whereas for PFOA total antioxidant capacity peaked at 1000 mg kg -1 of PFOA exposed in neutral and alkaline soils, while in OECD soil it was attained at early concentration i.e. approximately 400 mg kg -1. The antioxidant capacity began to decline at higher concentrations in both PFOS and PFOA. 100

101 Figure 6.5 Lipid peroxidation in earthworms exposed to (a) PFOS and (b) PFOA they are expressed in terms of MDA content 101

102 6.3.4 Lipid peroxidation In this study, the impact of PFOS and PFOA on lipid peroxidation in earthworms was studied. It is evident in Fig 6.5 a,b that the production of MDA was higher in treatments compared to control for both PFOS and PFOA exposure. In the case of PFOS exposure, the OECD soil showed a higher lipid peroxidation rate compared to the other 2 soils as observed by the MDA formation. The rate of peroxidation was higher in the OECD soil with maximum effect observed at 300 mg kg -1. With reference to the other two soils, lipid peroxidation which was measured in terms of MDA was higher only at 500 mg kg -1 when compared to the control. However, throughout the treatment we observed a dose dependent response in all 3 soils,; in fact lipid peroxidation is highly significant above 100 mg kg -1 of PFOS in all three soils. According to Fig 6.5 the production of MDA in PFOA treatment is less compared to the PFOS exposed organism. Treatment with PFOA and PFOS in the OECD soil appeared to strongly induce lipid peroxidation compared to the other 2 soils. The MDA content steadily rose from 50 mg kg -1 to 500 mg kg -1. Lipid peroxidation is significant above 100 mg kg -1 in the OECD soil for PFOA treated earthworms. By contrast, in the natural soil there was no significant response in MDA production up to 300 mg kg -1 of PFOA. Later there was a simultaneous increase in the concentration and duration of exposure and rate of lipid peroxidation. 6.4 Discussion Effect of PFOS and PFOA on E. fetida defense system Once an organism is exposed to stress condition, maintenance of equilibrium in reactive oxygen species (ROS), is disturbed to the extent that generation of ROS exceeds the scavenging capacity of the antioxidant enzymes. The drift in the defence system of the organism leads to various physiological malfunctions to most of the cellular compounds like proteins, DNA and lipids (Nel et al. 2006). An organism s defence system is triggered by the generation of antioxidant enzymes that combat the oxidative stress. Antioxidants play an important role in preventing the formation and scavenging of free radicals and other potentially toxic oxidising species. There are three antioxidant species; enzyme systems (GSH reductase, catalase, peroxidase, etc.); small molecules (ascorbate, uric acid, GSH, vitamin E, etc.); and proteins (albumin, transferrin, etc.). Of these small molecules, reduced glutathione (GSH) is an in vivo tripeptide that is critical in organising the defence system of an organism; it possesses an unusual linkage between the amine group of cysteine and 102

103 carboxyl group of the glutamate side chain. The detoxification process of GSH mainly concerns the bondage between the sulphydryl group with carbon or chlorine atom of the pollutants and their metabolites (Zelikoff et al. 1996; Xu, Dongmei et al. 2013b). Variation in the total antioxidant capacity in E. fetida is highlighted in Fig 6.4, and this variation is indirectly related to the depletion of GSH at higher concentrations. Once there is a reduction in GSH content, the pollutants and their metabolites escape the defence system s attack leading to the formation of larger biological molecules. The outcome of this is an oxidative stress condition. Even the study conducted using carp (Cyprinus carpio) reported the ability to generate oxidative stress, in turn leading to disturbance in the carp s DNA integrity (Hoff, P et al. 2003). One study conducted with tilapia hepatocytes reported that PFOA is able to elevate certain antioxidant enzymes like CAT, SOD and glutathione reductase (GR) at an initial stage (Liu, Chunsheng et al. 2007a). Excessive generation of ROS accompanied by increasing activity of superoxide dismutase (SOD), catalase (CAT) and glutathione reductase (GR) were detected, along with retardation in glutathione peroxidase (GPx) and glutathione-s-transferase (GST) following the degradation of glutathione (GSH) content on exposure to PFOA and PFOS in fresh water tilapia (Oreochromis niloticus) (Liu, Chunsheng et al. 2007a). Responses were similar in earthworms used in this study but the pattern of total antioxidant capacity expression differs from soil to soil and remains unclear. The OECD soil showed higher toxic effects for both PFOS and PFOA exposure, suggesting that exposure to perfluorinated compounds resulted in the production of ROS. This had the effect of stimulating the antioxidant system which further led to other physiological systems functioning poorly and the formation of genetic defects. Contamination leads to the induction of antioxidant enzymes which assists the organism to overcome stress that has resulted from exposure to an unhealthy situation (Cossu et al. 1997). However, unlimited stress hinders the normal functioning of these enzymes. Normally, ROS levels are in equilibrium with the amount of antioxidant enzymes. However, excessive generation of ROS either directly or indirectly by anthropogenic contaminants cannot be cleared correctly, so the natural antioxidant defences can be impaired. This results in fatal sub-cellular damage such as ion loss, protein denaturation, and DNA impairment (Nel et al. 2006). PFOA on chronic exposure induce excessive ROS generation in green algae leading to cellular injuries (Xu, Dongmei et al. 2013a). This finding supports the present study where PFOA induced oxidative damage in earthworms. 103

104 6.4.2 Lipid peroxidation induced by PFOS and PFOA in earthworms Lipid peroxidation, which is responsible for inducing various kinds of cell damage, is the most widely accepted marker of oxidative stress in the organism (Lushchak et al. 2001; Dalton, Puga & Shertzer 2002). Investigating lipid peroxidation is very important because ROS generation is directly proportional to the amount of fatty acid metabolism that occurs in the body. Lipid peroxidation in an organism is measured in terms of MDA, which is a product of oxidative damage. The level of free radicals in an organism is measured indirectly by the presence of MDA in the organism that causes cell damage. Our results indicated the formation of MDA in all the treatments exposed to both PFOS and PFOA. This was consistent with the study conducted using common carp, where MDA production increased due to PFOS and PFOA exposure mainly in the liver and kidney (Arukwe, Augustine & Mortensen, Anne S 2011). Thus our result indicated the generation of ROS, and activation of defence system in the organism. The complete process could be responsible for the DNA damage in the organism (Xu, Dongmei et al. 2013b). PFOS and PFOA concentration plays a vital role in affecting the biochemical responses identified inour study. Most in vitro and in vivo studies conducted using fish and mammals have suggested that ROS are highly responsible for metabolism of polyunsaturated fatty acids, namely lipid peroxidation which affects most of the body s physiological functions (Arukwe, Augustine & Mortensen, Anne S 2011). Interestingly, cultured hepatocytes of freshwater tilapia (Oreochromis niloticus) exposed to PFOS and PFOA resulted in the elevation of the lipid peroxidation (LPO) level (measured as maleic dialdehyde, MDA). This was noted only in the PFOA exposure groups, whereas LPO remained unaltered in the PFOS exposure groups (Liu, Chunsheng et al. 2007b). In contrast our study reported that lipid peroxidation increased in response to PFOS exposure rather than PFOA. Escalation in MDA content could be connected to the appearance of ROS. It has been reported that salmon continuously fed with dietary PFOS or PFOA dose displayed changes in peroxisomal responses and oxidative stress responses, with marked harshness in the kidney compared to liver (Arukwe, Augustine & Mortensen, Anne S. 2011). Advances in science have demonstrated that PFDoA leads to abnormal retention of fat in the liver due to failure in normal synthesis and elimination of fat in rats. This acts as a novel toxicological mechanism of PFCAs at the systems level (Ding, L et al. 2009). Support for this is given by studies on gene expression, where genes related to lipids transportation, inflammation and immunity and cell adhesion were suppressed in rat liver exposed to PFOA (Guruge, Keerthi S et al. 2006). 104

105 Overall, our data suggest that ROS-induced oxidative damage may be irresistible and PFOA and PFOS have a toxic impact on earthworms as has been suggested in aquatic animals DNA damage induced by PFOS and PFOA Our study suggested that the generation of excessive ROS leads to an oxidative stress state that ultimately impairs the DNA of E. fetida. Enough 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). An experiment conducted utilising common carp exposed to various concentrations of PFOS (0.1, 0.5 and 1 mg/l) resulted in the expression of various genes related to energy, reproduction and stress response after 14 days of exposure (Hagenaars, A et al. 2008). It has also been reported that PFOS induced stress that alters the expression of genes is related to the antioxidant system (Nakayama et al. 2008). According to one study conducted with paramecium, PFOA has a greater tendency to cause DNA damage than PFOS. Similarly, our study showed that % tail DNA initiation is higher in PFOA exposed organisms than in PFOS even though PFOS is highly toxic (Kawamoto, Kosuke et al. 2010). An investigation employing five different PFCs in generating reactive oxygen species (ROS) and oxidative DNA damage in HepG2 in humans, reported that perfluorooctanoic acid (PFOA) and perfluorooctane sulfonate (PFOS) increased the intracellular ROS production by 1.52-fold (95% CI, ) and 1.25-fold (95% CI, ), respectively. However, the increase in ROS production was not dose dependent and the compounds did not induce DNA damage that could be detected by the alkaline comet assay as strand breakage and alkali-labile sites or formamidopyrimidine- DNA-glycosylase (FPG) sites (Eriksen, Kirsten Thorup et al. 2010). It can be concluded that PFCs are effective in inducing DNA damage in lower organisms like earthworms rather than higher animals like humans. Our findings are consistent with those of a previous study in which PFCs activated the stress responsive genes which are triggered by DNA damage along with oxidative damage, general cell lesions and membrane damage in E. coli (Nobels et al. 2010). Hoff et al. (2003) concluded that PFOS could alter the average DNA base-pair length in C. carpio, speculating that PFOS obstruct homeostasis of the DNA metabolism (Hoff, P et al. 2003). Investigations of PFC have shown that PFOA treatment produced vitellogenesis and antioxidative tension in the common carp, C. carpio, while PFOS generated DNA damage. Hoff et al. (2003) added that biological responses should be discussed in ecological risk assessments of PFCs in fish. Similarly, a risk assessment concerning a terrestrial ecosystem should address the biological responses obtained from earthworms. In the case of 105

106 higher organisms like humans, PFOA and PFOS exert a cytotoxic effect on the human cell line HepG2 but PFOA and PFOS could not cause more DNA damage (i.e. DNA strand breaks and micronucleus) or reactive oxygen species at the tested concentration range. Hoff et al. (2003) also indicated that oxidative stress and DNA damage are not correlated to exposure of PFOA and PFOS. This finding encouraged epidemiological studies that did not document tumour induction (Eriksen, Kirsten T et al. 2009; Florentin et al. 2011). This does not apply to earthworms because in our study we clearly observed oxidative DNA damage. 6.5 Conclusion The present study demonstrated that acute PFOS and PFOA exposure could overwhelm the balance in antioxidative systems, resulting in oxidative stress that could lead to lipid peroxidation and thus DNA damage. The results produced important information that at least in part addresses an alarming concern for ecological risk in response to acute exposure to PFOS and PFOA. 106

107 7 Accumulation of PFOS and Behavioural Effects 7.1 Introduction Perfluorinated compounds are an important class of global organic contaminant and they are widespread in the natural environment (Paul, Jones & Sweetman 2008). These compounds have had wide industrial application such as fabric protectant, firefighting foam, oil resistant coatings, insecticides and even as surfactants (Kissa 2001; Houde et al. 2006). A newly emerging persistent organic pollutant, PFOS has been the subject of recent investigation because they can bioaccumulate especially at higher tropic levels (Houde et al. 2008). A report on the global distribution of these perfluorinated compounds indicated that concentration of PFOS is higher in samples from industrialised areas like the North American Great Lakes, Baltic Sea, and Mediterranean Sea rather than from the Arctic and Pacific regions (Giesy, JP & Kannan 2001). Due to the widespread use of these chemicals in in the last decade they are now distributed globally, and subsequently these compounds are detected in birds, animals and people (So, M et al. 2004; Hoff et al. 2005; Delinsky et al. 2010; Schecter et al. 2010). PFOS is highly bioaccumulative in nature and cable of inducing adverse conditions such as toxicity in liver, immunotoxicity, reproductive damage, developmental disturbances, carcinogenic-inducing tumour in pancreas, liver and breast in living organisms according to various toxicological studies using laboratory animals (Harada, Kouji et al. 2005; Cui et al. 2009). Unlike other POPs which are mostly lipophilic, these perfluorinated compounds showed much affinity towards blood protein and ultimately end up in the livers and gall bladders of organisms (Renner, Rebecca 2001). One elaborate review on perfluorinated compounds suggested acute and chronic toxicity of PFOS, PFOA and 8:2 FTOH ranged between moderate to low in aquatic organisms (Hekster, Laane & de Voogt 2003) and also indicated the presence of PFAS above the detection limit in all biota and human beings sampled from contaminated and non-contaminated sites. 107

108 Soils act as a valuable source for the ecosystem and promote the activities of soildwelling organisms in order to maintain its proper equilibrium. Therefore in order to develop a complex toxicity profile for PFOS acute toxicity in earthworms should be investigated. Currently, many societies agricultural soils are highly contaminated with perfluorinated compounds like PFOS, PFOA (Renner, Rebecca 2008). Prevalent sentinel organisms employed to study the toxicity of PFOS and PFOA are aquatic midge, fish, snails and plants (MacDonald, Michelle M. et al. 2004; Li, MH 2009; Dorts et al. 2011). Toxicity of PFOS varies from organism to organism. For instance PFOS is less toxic to onions whereas it is highly lethal in soybean observed from 21d EC 50 values. This type of variation is mainly due to the bioavailability of certain substances to particular organisms (Hirano & Tamae 2011). Earthworms are critical in the formation of soil because they decompose organic matter and their movement facilitates the mixing of fragmented organic matter with soil particles intensively, resulting in the formation of water stable aggregates (Edwards 2004a). The ideal organism for biomonitoring contaminated sites is the earthworm species E. fetida, which is the common sentinel organism employed in most toxicological studies, because these organisms are more closely linked to soil contaminant than other organisms (Reinecke, AJ & Reinecke, SA 2004; Steenbergen et al. 2005). Their advantage in biomonitoring of soil pollution is their ability to bioaccumulate contaminants (Hirano & Tamae 2011). While there are numerous chemical assays to estimate the toxicity of chemicals, much more attention is needed on biological responses because these help to understand the mechanism of toxicity in the organism. Sensitivity of ecotoxicological studies varies between laboratory and field conditions (Van Gestel, CA 1997). Most toxicological research has been conducted on laboratory animals only under controlled conditions (Saint-Denis et al. 2001; Reinecke, SA, Helling & Reinecke 2002). Field conditions are highly solitary in nature and often associated with complex parameters, so any toxic responses under such conditions act as a valuable biomarker that closes the gap between the results obtained in the lab and field (Spurgeon, David J, Weeks & Van Gestel 2003). Popular biomarkers used to assess the toxicity of contaminants under such unique conditions found in the field are responses from the sub-organism level. Sub-cellular responses are close related to the endpoints at the higher organism level (Reinecke, SA, Helling & Reinecke 2002). Currently, there are numerous discrepancies in the ecotoxicity test used for earthworm species of higher ecological relevance and whole endpoints that are directly related to ecological function in the soil (Capowiez et al. 2010). Thus we decided to investigate the toxicity of PFOS by recording the biological responses at sub-cellular level and link them 108

109 with behavioural changes such as alteration in locomotion, cast production and wound healing capacity. Two cytotoxicity assays were carried out: (1) LDH cytotoxicity, and (2) MTT assay. We also studied the bioaccumulative phenomena of PFOS in the earthworms using 3 different soils. 7.2 Materials and Methods Refer chapter Results Figure 7.1 Bioaccumulation factors of PFOS in earthworms exposed in 3 different soils Bioaccumulation of PFOS The mean PFOS concentrations in the whole composite body homogenates of earthworms were represented by BAF (bioaccumulation factor). PFOS was detected in all the treated organisms as shown in Fig 7.1. This indicated that PFOS are highly bioaccumulative in nature. The magnitude of contamination varied between the soils. No main difference was observed in the PFOS pattern between neutral and alkaline soil which showed the maximum BCF values of 2.12 and 2.23 at 1 mg kg -1, respectively. In fact the OECD soil showed the highest BCF value and this was approximately 3.8 at a lower concentration of around 1 mg kg -1. The pattern of bioaccumulation decreased as the concentration rose. The analysis of PFOS in earthworms exposed in 3 different soils showed that this compound is predominant in all treatment samples. A higher proportion was detected 109

110 in the OECD soil samples. We could conclude that the PFOS accumulate more at smaller concentrations of 1-50 mg kg -1. These concentrations were reported to have a higher BCF factor. Higher concentrations such as mg kg -1 all reported having a BCF factor of between Cytotoxicity of PFOS The result of LDH cytotoxicity on Eisenia fetida exposed in 3 different soils (neutral, alkaline and OECD soil) to different concentrations of PFOS are shown are shown in Fig 7.2. One way ANOVA was conducted to identify the significance of effect in the treatments. A significant dose response relationship was observed in all the treatments in all 3 soil types. Dunnett t test (2-sided) was conducted to observe the significant effect of the treatment to that of the control. Concentrations above 100 mg kg -1 were significant (p<0.05) in impacting on the OECD soil, neutral soil. and alkaline soil. The percentage of cytotoxicity observed in all the controls was around 8.1%. From It is evident in Fig 7.2 that the behaviour of PFOS in inducing cytotoxicity was similar in all 3 soils up to 200 mg kg -1. Later the OECD soil showed higher cell toxicity compared to the natural soil. Intensity of inducing cell death was higher in OECD soil compared to the natural soils with complete cell death occurring at maximum concentration exposure. The results of MTT assay conducted on Eisenis fetida exposed to different concentrations of PFOS in neutral, alkaline and OECD soils are shown in Fig 7.3. The percentage of MTT reduction in earthworms exposed to PFOS was calculated in terms of vitality (percentage of control). The reduction of MTT in the control earthworms, was set at 100%. Cell vitality decreased when the exposure concentration increased. The pattern of cell toxicity was similar in all 3 soils. Surprisingly the effect on cells is stronger even at 100 mg kg -1 only in neutral soil and more signfiicant than the control (p<0.05) using Dunnetts t test (2-sided). However, in the OECD soil and alkaline soil a significant effect was observed only after 200 mg kg -1. Importantly the toxic effect of PFOS was stronger in OECD soil soon after 100 mg kg -1. Viablity was reduced to < 20% at maximum exposure dose in OECD soil 300 mg kg -1 and in natural soils at 500 mg kg

111 Figure 7.2 Cytotoxicity (%) observed in earthworm treated with different concentration of PFOS maintained in 3 different soils (neutral, alkaline and OECD). Reduction of the colorant MTT by coelomic cells of Eisenia fetida after exposure for 14 days to different concentrations of PFOS; cell vitality was determined as the mean of the absorption at 570 nm and expressed as %± SD of the control. Cell vitality was determined as the mean of the absorption at 450 nm and expressed as %± SD of the control. 111

112 7.3.3 Locomotion Locomotor ability was measured on day 14 after acute exposure to PFOS. Crawled distance in one minute is shown in Fig. 4. The mobility was affected by PFOS at the earlier concentration of 10 mg kg -1 in OECD soil and 50 mg kg -1 in neutral and alkaline soils. Although the locomotor ability of earthworms exposed to a range of soil PFOS concentrations generally declined when PFOS concentrations rose, there was no difference between the natural soils, but the OECD soil had more impact on earthworms mobility. Significant dose response relationship was observed in all the treatments. In OECD soil the effect on mobility when exposed to PFOS was significant even at a smaller concentration 10 mg kg -1 (p<0.05). In natural soils locomotion was affected only at 50 mg kg -1 and highly significant to that of the control (p<0.05). The significance was calculated using Dunnett t test (2-sided) with One way ANOVA. Results indicated that PFOS is strong in affecting their locomotion ability.. Figure 7.4 Mean (± SD) crawled distance by earthworm in one minute on 14 day after exposing to different PFOS concentration maintained in 3 different soils. 112

113 Figure 7.5 Percent of earthworms having complete healing after 5 d post wounding on 7, 14 and 28-d exposure in 3 different soils after treating with PFOS. Sample size is 20 for each group. 113

114 7.3.4 Wound healing Capacity Results of the 5-d wound healing capacity are expressed as healing percentage according to the concentration and duration of exposure (see Fig 7.5). The sensitivity of wound healing capacity is highly responsive to exposure concentration and duration. Both the concentration and exposure significantly affected the wound healing capacity of earthworms exposed in 3 different soils. In Fig. 5 we can clearly see the effect of PFOS varying significantly from the 7 th day, 14 th day and 28 th day. PFOS effect on the wound healing capacity of earthworms exhibited different patterns, even natural soil showed a variation in their effect. In OECD soil the dose response relationship was significant at 100 mg kg -1 on the 7 th day of exposure, but on 14 th day and 28 th day the effect was stronger which showed a significant effect on lower concentration of 10 mg kg -1. In the case of neutral soil PFOS induced a significant effect on wound healing capacity at 100 mg kg -1, and the extended exposure duration at 14 th and 28 th days resulted in a significant effect even at a lower concentration of 50 mg kg -1. The effect was less or the alkaline soil and any significant affect was only observed at 200 mg kg-1 on the 7 th day. However, the effect increased at a lower rate on the 14 th day and 28 th at 100 mg kg -1. It is evident that the effect of PFOS on wound healing is highly sensitive to dose and exposure duration. Figure 7.6 Effect of different concentrations of PFOS on earthworm cast production (means and standard deviations) expressed in % of soil available (each earthworm was given 300 g of moist soil) expressed in g/worm/day. Cast production was estimated immediately at the end of the 14 day exposure period (no soil drying). 114

115 7.3.5 Cast production Cast production was difficult to determine using the OECD soil, because the cast could not be easily identified due to the soil condition. We conducted a cast production test on natural soils (neutral and alkaline soil). Sieving was carried out immediately after 14 days exposure. PFOS had a significant effect on the both neutral and alkaline soils. Significant dose reponse relationship was exihibited in all treatments. Dunnett t test (2-sided) helped to identify the concentration that had a significant effect. Both soils at a low concentration of 50 mg kg -1 highly affected the cast production in the earthworms (p<0.001). A 50% reduction in cast production was observed earlier in the neutral soil at 100 mg kg -1 but in the alkaline soil the effect causedd 50% reduction in cast production at only 200 mg kg -1. We can conclude that cast production acts as a sensitive indicator for the effect of PFOS on acute exposure. 7.4 Discussion Bioaccumulation of PFOS Recently, much attention has been focused on PFOS because of their bioaccumulation capacity. Despite relatively higher soil concentrations of PFOS, the body burden of PFOS in earthworms at the end of the 14-day experiment reached stauration. The data presented in Fig 7.1 suggest that PFOS reached a steady state condition after 14 days under laboratory conditions. Predominantly the lower concentrations of PFOS are relatively higher in accumulation especially in OECD soil compared to the natural soils. Bioaccumulation of PFOS is determined by calculating the biaacumulation factor (BAF) in earthworms after exposing them to different concentrations of PFOS. BAF values at higher concentrations of PFOS remains constant. PFOS bioaccumulated in the earthworms as shown in Fig 7.1, which also suggests that in lower concentrations the BAF value are more for OECD, neutral and alkaline soils. OECD soil showed the highest BAF value for PFOS. An earlier study indicated that PFOS are highly bioaccumulative in nature, and BAF value was estimated to be 3.4 in lake trout (Houde et al. 2008). PFCs with longer chains are highly detectable in biota, whereas shorter chain compounds are less accumulative in rainbow trout (Martin, Jonathan W et al. 2003). Examination of PFC in Lumbriculus variegatus indicated that biota sediment accumulation factor for PFOS was higher compared to other PFCs due to more accumulation in the body. The pattern of bioaacumulation tended to decrease as the chain length decreased. However, bioaccumulation of PFOS is unclear and one study adviced that PFOS has lower elimination constant than ther PFCs. This reason might be responsible for higher bioaccumation in biota (Higgins et al. 2007). Our study indicaed much variation in 115

116 bioaccumlation patterns, which could be concluded from the BAF value obtained in Fig Previous reports indicated that soil properties influence more the bioavailability of PFOS and PFOA. Phytotoxicity study of perfluorinated compounds on Brassica chinensis postulated that soil with high organic matter significantly alters the bioavailability of PFOS and PFOA and also suggested that perfluorinated compounds created different toxicities in plants depending on the soil properties (Zhao, H et al. 2011). Unlike other chlorinated organic pollutants like dioxin which are mainly fat-loving compounds and highly detectable in cecca and depot fat in rainbow trout (Muir, D et al. 1986), PFAs are not lipophilic. Instead they showed more affinity to plasma albumins (Guy, Taves & Brey Jr 1976). The general assumption was that PFSAs target mostly the proteins produced and secreted by the liver such as carrier proteins for fatty acids (Vanden Heuvel et al. 1991). Earthworms are enriched by albumin so the accumulation of PFOS in our experiment is higher because of their affinity to these albumins. There are previous information on accumulation of long chain perfluorinated acids in living samples from the Canadian Arctic this highlights the efficiency of PFCs in bioaccumulation (Martin, J. W. et al. 2004). Our report suggests that PFOS is highly accumulative in the earthworm tissues Effect on cell viability The LDH cytotoxicity assay and MTT were conducetd photometrically in earthworms coloemocytes to evaluate the toxicity of PFOS. Our study demonstrated that the two assays helped to assess the cytotoxic capacity of PFOS which induce cell death in all treatments (see Fig 7.2 and 7.3). Cell death was evaluated by plasma membrane damage. Lactase dehydragenase is a stable enzyme present in all types of cells and rapidly released when there is damage to the plasma membrane. Our results indicated that all the treatments induced cell death in earthworms after exposing to PFOS. Significant effect on plasma membrane damage was observed at 100 mg kg -1 in neutral, OECD and alkaline soils. Figure 7.2 illustrates that the release of LDH enzyme is stronger at higher doses of PFOS due to their lower metabolic activity. Significantly less redution in MTT dye was observed at 100 mg kg -1 in neutral soil, thus indicating less metabolic activity whereas in the alkaline and OECD soils it was observed only at 200 mg kg -1. A similar effect was observed in earthworms coelomocytes regarding long-term exposure to cadmium (Maleri et al. 2008). Dose effect relationship was exihibited in both the assay in all the treatments exposed to PFOS. Cytotoxic effect in coelomic fluid of Eisenia foetida has been previously evaluated by observing the destruction of cell membranes. Antibacterial effect was observed in 116

117 earthworms using LDH release (Kauschke & Mohrig 1987). Acute exposure to PFOS produced a significant response in both in MTT assay and LDH cytotoxicity assay. This indicated that sub-cellular damage is prominent in earthworms exposed to higher concentrations of PFOS. Most studies on cell cytotoxcity using MTT were restricted only to metal comtamination so there are no comparable data for our study. PFOS disturbs the defence system of the earthworms to a great extent, an earlier study revealed that normal functioning of the antoxidant system was interrupted by PFOS exposure. A reduction in SOD POD activity occurred due to excessive generation of ROS in the system (Xu, Dongmei et al. 2013b). A balance in ROS generation and antioxidant system is needed for normal life functioning. However, excess ROS produced during xenobioc metabolism cant be scavenged with normal antioxidant system. This is accountable for many sub-cellular injuries such as ion loss, protein denaturation and DNA damage (Nel et al. 2006). Thus the production of ROS might be responsible for cell death as seen in our current study. The two photometric assays - MTT and LDH cytotoxicity assay - were highly useful in evaluating the toxic effects of PFOS at the sub-cellular level Effect on Locomotion ability Earthworms have the capacity to avoid harsh environments. They clearly avoid even low contamination levels before anticipating negative effects that are estimated through acute toxicity test and reproduction test. Avoidance can be exhibited in many ways, and they include the ability to locomote inside the stress condition and be highly sensitive to the contaminant dose (Zheng, R & Li 2009). Our findings showed that as the PFOS concentration increased the mobility of the organism was much reduced. It can be speculated that the mobility of earthworms was much influenced by PFOS. One analysis of locomotion looked at the influence of lead concentration in earthworms; it had a very large effect at higher concentrations. The test had some drawbacks such as locomotion ability can also be affected by other factors such as other contaminants. We compared the results to that of the control and concluded that the effect might be due to stress caused by PFOS. It has been documented that PFOS induced significant avoidance behaviour in earthworms at 160 mg kg -1 (Xu, D et al. 2011) and that avoidance is more sensitive in natural soil compared to artificial soil. Our results also indicated that PFOS is significant even at 10 mg kg -1 in affecting the mobility of earthworms under artificial soil, but in natural soil at 50 mg kg -1 it is predicted that OECD soil is more sensitive than neutral soil. PFOS is reported be developmental neurotoxicant that produce an abnormal behavioural response in 2- and 4-month old mice exposured to PFOS 117

118 (0.75 or 11.3 mg kg -1 body weight) (Johansson, Niclas, Fredriksson, A & Eriksson, Per 2008). It can therefore be speculated that PFOS could infect the nervous system of earthworms and thus produce abnormal locomotion patterns Effect on wound healing capacity The wound healing capacity of earthworms appeared to be highly responsive to both concentration and exposure. Our results reported that both concentration and duration of exposure significantly hindered the process of wound healing in earthworms. Worms showed reduced wound healing capacity at all exposed concentrations, but the effect was stronger in OECD soil where 10 mg kg -1 caused a significant reduction compared to natural soils. Earlier studies postulated that wound healing capacity in earthworms was suppressed by chlordane, and some organics such as PCB, pentachlorophenol (PCP), inhibited phagocytosis in earthworms coelomocytes, probably affecting cell membranes (Goven et al. 1994; Giggleman et al. 1998; Cikutovic, MA et al. 1999). Tissue repair sytem in an organism is very rapid and mediated by Ca 2+ influx precipitated by disruption in the plasma membrane (Reddy, Caler & Andrews 2001). Thus the maintenance of plasma membrane from loss along with restoration ensures viability and longevitity of cells (McNeil, P. L. & Steinhardt 1997). The current study showed that PFOS actively ruptured the plamsa membrane estimated by LDH release (see Fig. 7.2). Another possible reason for inhibition of would healing capacity is damage in RNA synthesis which could interfere with the tissue repair system of the organism. This was proved by one study conducted in earthworms exposed to lindane and other chlorinated pesticides (Thomas PT 1895). Any alteration in cell division may compromise wound healing ability. Some compounds like chlordane and cadmium disturb spermatogenesis in L. terrestris by changing cell division (Cikutovic, M et al. 1993). Some heavy metals also generate similar effects on the tissue repair system. Cadmium encroaching on the antioxidant system of the earthworms leads to immune deficiency (Chen et al. 1991). Similarly, PFOS interfere in the antioxidant system of earthworms, and reports have suggested that initial concentrations of PFOS slightly initiated antioxidant functioning but later stopped them completely (Xu, Dongmei et al. 2013b). These authors also reported that PFOS is genotoxic. Although it is very difficult to understand the process of interaction of PFOS on the wound healing capacity, the effects on cell division, antioxidant system, DNA / RNA sysnthesis and energy production may be enough to suppress the proper funtioning of the tissue repair system (Cikutovic, MA et al. 1999). Its malfunctioning may help to forecast the damage induced by environmental toxicants well in advance if it is wounded naturally. 118

119 Thus any natural wound such as punctures, lacerations, and abrasions may mean vulnerabilityto pathogens if an environmental toxicant suppresses the wound healing As a biomarker wound healing capacity has potential in determining the effect of PFOS in a terretrial ecosystem Effect on Cast production In our study cast production helped to identify the effect of PFOS on acute esposure. Reduced cast generation in the presence of PFOS appears to be a sensitive response to toxicants. Most of the concentrations showed a highly significant response in cast production as shown in Fig. 6. This effect was similar in neutral and alkaline soils even there was some variation between these two soils. An earlier study conducted on earthworm (Lumbricus terrestris) to determine the effect of 6 pesticides ((imidacloprid, carbaryl, methomyl, ethylparathion and chlorpyrifos-ethyl) on cast production reported similar results. They showed a significant dose effect relationship that correlated the effect on cast production with weight loss (Capowiez et al. 2010). Cast production is highly sensitive because it has more ecological value than other responses. We can see that the effect on cast production is sensitive at lower concentrations compared to other sub-cellular responses. Previously, it was reported that PFOS is toxic to terrestrial inverterbates. The LC 50 value for PFOS in 14 day acute toxcity assay was 365 mg kg -1 which was also significant in reducing body mass occurred (Joung et al. 2010). It can therefore be concluded that even the concentrations below LC 50 are significant in inducing behavioural responses such as reduced cast production. It was found that PFOS induce significant variations in growth on acute exposure in earthworms, and suggested that alterations in growth may affect functions at the molecular and physiological levels (Xu, Dongmei et al. 2013c). Our study showed that PFOS is effective in reducing the mobility of earthworms (see Fig. 7. 4). It can be linked to reduced cast production because they fail to burrow inside the soil, remain docile and ingest PFOS conatminated soil. 7.5 Conclusion From our experimental study we conclude that PFOS is highly accumulative in earthworm tissues. They are also able to induce cellular damage. Most of the ecological parameters in earthworms were susceptible to PFOS burden. 119

120 8 Accumulation of PFOA and Behavioural effects 8.1 Introduction Many reviews have documented the fate and behaviour of PFCAs, PFSAs, and related PFSs in the environment. The presence of PFCs in biota is inevitable, and it has been reported that PFSs can attain maximum concentrations ng ml -1 or ng g -1 of wet weight (ww) in humans and wildlife dwelling in urbanised areas of North America, Europe, and Asia. PFSs also reach elevated levels ( 3000 ng g -1 ww) in organisms ranging from the Arctic to the ocean do islands (Houde et al. 2006). Most of the PFCs are manufactured synthetically and used as common ingredients in commercially available products, which are also released accidently as by-products. Unfortunately, there is presently little information on the chemical physical properties of most PFCs, and even less toxicity information is available on these compounds (Giesy, J et al. 2010). Perfluorooctanoic acid (PFOA) is a member of the perfluoroalkyl acids that have many industrial applications. It is also a pervasive pollutant having a significant toxicological impact noted in environmental matrices (Son et al. 2008). PFOA is widely distributed such that its presence is invariable in both exposed and non-exposed organisms. The semi-field study conducted on general human populations detected low level of PFOS in their serum ( ppb) whereas occupationally exposed workers had higher concentrations of PFOA. Despite data on the presence of PFOA in humans there is no much information on its toxic effect on humans. One study on rodents reported that PFOA is a potential hepatocarcinogen which also alters the reproductive hormones in both humans and rodents (Gilliland & Mandel 1993). PFOA is highly lethal in rodents and induces death within 5 days. It was reported from the same study the LC 50 value for PFOA was around 189 mg kg -1 (range ) upon 30-day exposure. Some toxic effects have been identified like significant increase in liver weight associated with reduced food intake. They also promoted increased fatty acid activities in oleic and palmitic acid but reduced stearic and docosahexaenoic acid in rodents (Olson & Andersen 1983). A more 120

121 recent acute toxicity study on mice reported that PFOA affects the liver but not the kidney. The organism became inactive at 250 mg kg -1 with a significant weight reduction. PFOA alters the serum enzyme activities, alanine aminotransferase and aspartate aminotransferase in mice (Son et al. 2008). PFOA is less bioaccumulative compared to PFOS. An experimental study on one day-old male chickens indicated that PFOS and PFDA accumulated at much higher concentration than PFOA. It also measured the half-lives for each PFC at 0.1 and 1.0 mg kg -1 concentrations that were, respectively, approximately 15 and 17 days for PFOS, 11 and 16 days for PFDA, and 3.9 and 3.9 days for PFOA. The BCF of PFOS in fish were measured to be about 10,000 and above, while the BMF of PFOS in fisheating birds and minks were in the range. However, the BCF of PFOA in fish was only about 200 or less, much lesser than that of PFOS (Fujii et al. 2007). Even though PFOA are less likely to bioaccumulate they impose greater risks for mortality from prostate cancer in workers exposed to PFOA (Gilliland & Mandel 1993). PFOA creates significant hazards for plants like Brassica chinensis, and one study reported the EC 50 value for PFOA ranged from 107 to 246 mg kg -1. The effect also varies depending upon the soil parameters especially organic matter (Zhao, H et al. 2011). A recent study determined significant, concentrationdependent carry-over of PFOA and PFOS in crop plants, as acting as a potential entrance point for these substances into the food chain (Stahl et al. 2009). Terrestrial organisms come in contact with these chemicals mainly through food, drinking water and atmospheric exposure. The main organism employed for biomonitoring contaminated sites is the earthworm. The earthworm species (E. fetida) is the common bioindicator used in most toxicological studies, because it comes into frequent contact with soil contaminants than any other organism (Reinecke, AJ & Reinecke, SA 2004; Steenbergen et al. 2005). These organisms are capable of bioaccumulating the xenobiotics that are highly beneficial in biomonitoring contaminated sites (Hirano & Tamae 2011). This study investigates the multiple effects of PFOA in earthworms (Eisenia fetida) and also determines the link between various biological responses by identifying the sub-cellular responses and their physiological endpoints. 8.2 Materials and methods: Refer chapter Results 121

122 Figure 8.1 Bioaccumulation factor of PFOA in earthworms exposed in 3 different soils (neutral, alkaline and OECD soils) Bioaccumulation of PFOA in earthworms The mean PFOA concentration in the whole composite body homogenates of earthworms were represented by BAF (bioaccumulation factor). PFOA was detected in all treatments (Fig 8.1). This indicates that PFOA accumulates in the earthworm s body. The magnitude of accumulation of PFOA varied between the 3 soils. No significant difference was observed in the PFOA accumulation pattern between natural soils with a maximum BCF value of 1.29 and 1.39 at 1 mg kg -1, respectively. Contrasting this, the OECD soil showed the highest BCF value at around 2, this being the lowest concentration (1 mg kg -1 ). The pattern of bioaccumulation decreased as the dose increased invariably in all soil types. The analysis of PFOA in earthworms exposed in 3 different soils showed that this compound is prevalently distributed in all treatment samples. The highest concentration was detected in earthworms in the OECD soil. We can conclude that PFOA appears to accumulate more at smaller concentrations from 1-50 mg kg -1, because these concentrations have a higher BCF factor (see Fig 8 1). A higher concentration such as mg kg -1 had a BCF factor of around

123 Figure 8.2 Cytotoxicity (%) observed in earthworm treated with different concentration of PFOA maintained in 3 different soils (neutral, alkaline and OECD). Reduction of the colorant MTT by coelomic cells of earthworms after exposure for 14 days to different concentrations of PFOA; cell vitality was determined as the mean of the absorption at 570 nm and expressed as %± SD of the control. Cell vitality was determined as the mean of the absorption at 450 nm and expressed as %± SD of the control. 123

124 8.3.2 Cytotoxicity of PFOA a) - The result of LDH cytotoxicity on Eisenia fetida exposed in 3 different soils (Neutral, Alkaline and OECD soil to different concentrations of PFOA - see (Fig. 8.2). One-way ANOVA identified the significance in the treatments to control using Dunett t test (2-sided). A significant dose response relationship was observed in all the treatments at higher concentrations in all 3 soil types. Concentrations above 300 mg kg -1 were significant (p<0.05) in effecting the OECD soil. whereas in neutral soil and alkaline soil a significant effect was observed at 500 mg kg -1. The percentage of cytotoxicity observed in all the controls were around 8.1%. From Fig. 2 we can observe that the pattern of inducing cytototoxicity was similar until 300 mmg kg -1 was reached in the 3 soils, as the concentration rise in the OECD soil appeared to be more toxic. Interestingly, in the natural soils mg kg -1 actually showed a reduction in cytotoxicty but as the dose increased the cytotxicity also rose. Intensity of inducing cell death was higher in OECD soil compared to natural soils with complete cell death occurring only at highest concentration exposure. (b) - The results of MTT assay conducted on Eisenia fetida exposed to different concentrations of PFOA in neutral, alkaline and OECD soil are shown in Fig. 8.2b. The percentage of MTT reduction in earthworms exposed to PFOA was calculated in terms of vitality (%) over that of control. The reduction of MTT in the control earthworms was set at 100%. Cell vitality dropped as PFOA concentration increased. The patttern of cell death was similar in all 3 soils. A clear dose response relationship was noted in the pattern of cell death in all soil types. When compared the effect of PFOA within the soil types the OECD soil was more toxic than other two.samples showed a significant effect at 200 mg kg -1, whereas natural soils induced a significant toxicity only above 300 mg kg -1. However, the alkaline soil was moderate in inducing this toxic effect in the treatment. 124

125 Figure 8.3 (a) Mean (±SD) crawled distance by earthworm in one minute on 14 day after exposing to different PFOA concentration in OECD soil (b) Mean distance travelled in one minute on day 7 and mean (±SD) of cast production (g/worm/day) on day 14 in earthworms exposed to PFOA in neutral soil, (c) Mean distance travelled in one minute on day 7 and mean (±SD) of cast production (g/worm/day) on day 14 in earthworms exposed to PFOA in alkaline soil. 125

126 8.3.4 Physiological end points Locomotion ability of earthworms was measured on day 14 after acute exposure to PFOA. Distance covered by the organism in one minute is shown in Fig 8.3 a,b,c in the OECD, neutral and alkaline soils, respectively. Although the locomotion of earthworms exposed to a range of soil PFOA concentrations generally declined as PFOA concentrations increased, there was no difference between the effect with in natural soils, but OECD soil evidenced a greater effect on the locomotion of earthworms. Significant dose response relationship was observed in all the treatments. The mobility in earthworms was affected by PFOA at the earlier concentration of 100 mg kg -1 in OECD soil and 200 mg kg-1 in natural soils. However, in OECD soil the effect on mobility upon PFOA exposure was significant only after 200 mg kg -1 (p<0.05). In natural soils (both alkaline and neutral soil) locomotion was affected only at 300 mg kg -1, which was significant compared to the control (p<0.05). This significance was calculated using the Dunnett t test (2-sided) with one way ANOVA. From the analysis the results indicated that PFOA is moderately toxic in affecting the earthworms locomotion. Cast production was difficult to determine using the OECD soil, because it was very difficult to identify the cast due to the soil condition. We carried out a cast production test on the natural soils (neutral and alkaline soil). Sieving was carried out immediately after 14 days exposure. PFOA induced a significant effect on both the neutral and alkaline soils. Significant concentration reponse relationship was exihibited in all treatments. Dunette t test (2-sided) was carried out to identify the concentration that generated a significant effect. There are some variations in the effect in the soil but they were not significant. A 50% reduction in cast production was observed at 300 mg kg -1 in neutral soil and alkaline soil. Even though a significant effect was observed above 300 mg kg -1 in the natural soil, this was more than that detectedin the alkaline soil. Thus we can conclude that cast production acts as a good bioindicator of the effect of PFOA on acute exposure. 126

127 Figure 8.4 Percent of earthworms having complete healing after 5 d post wounding on 7 th, 14 th and 28 th d exposure in 3 different soils treated with PFOA. Sample size is 20 for each group. 127

128 Results of the 5 day wound healing capacity are expressed as percentage of healing according to the concentration and duration of exposure (Fig. 8.4). The sensitivity of wound healing capacity is moderately responsive to exposure concnetration and duration. Both the concentration and exposure significantly affected the wound healing capacity of earthworms exposed in 3 different soils. In Fig. 5 we can clearly see the effect of PFOA and it variers significantly from the 7 th day, 14 th day and 28 th day. In OECD soil the dose response relationship was significant only at 300 mg kg-1 on the 7 th day of exposure, and 14 th day but on the 28 th day the effect was stronger, indicating a significant effect at a smaller concnetration of 200 mg kg -1. In the case of neutral soil PFOA induced a significant effect on the wound healing capacity at only 700 mg kg -1 whereas the extended exposure duration at 14 th day and 28 th day resulted in a significant effect even at a smaller concentration of 500 mg kg-1 on the 14 th day and 300 mg kg -1 at the 28 th day. The effect on wound healing capacity was similar in the alkaline soil. We can conclude that wound capacity is affected by PFOA but it is not as sensitive as that of subcelluar responses. 8.4 DISCUSSION Bioaccumulation of PFOA in earthworm The extent of variation in PFOA bioaccumulation among the 3 soils is highly significant in OECD soil compared to natural soils (see Fig. 8.1). Here the trend of accumlation increased as the dose of PFOA increased, but the BAF value was higher at lower concentrations. Similarly, a previous study conducted on green mussels, Perna viridis exposed to 1-10µg L -1 of PFCs for 56 days reported the BAF values for different PFCs ranging from 15 to 859 L Kg -1 and 12 to 473 L Kg -1. It also suggested that a lower dose resulted in higher BAF values. The same study suggested that the non-linear adsorption mechanism is responsible for the pattern of increased accumulation with respect to elevated dose exposure (Liu, C et al. 2011). Our study estimated the BAF value was 2 at a lower concentration of PFOA in OECD soil, whereas the neutral soil accumulation was 1.29 and 1.39 in terms of BAF for alkaline soil at the same concentration (Fig 8.1). The BAF in OECD soil ranged from , neutral soil , alkaline soil The bioconcentration of PFC with respect to their fluorinated chain length and functional group was observed in mussels (Liu, C et al. 2011), common carp (Inoue et al. 2012), Lumbricus variegates (Lasier et al. 2011), rainbow trout (Martin, Jonathan W et al. 2003) and even in higher mammals like the rat (Kudo et al. 2001). 128

129 BAF is highly sensitive to perfluorinated chain length and their ability to attract proteins. Though our study reported the bioaccumlation pattern of PFOA, general available research suggested that PFASs are more bioaccumulative than PFCAs with the same flourinated length. It also reported that PFCAs with seven carbon or less are not considered to be biaccumlative according to the range of bioaccumalation B regulatory criteria of L Kg -1 (Conder et al. 2008). However, we reported that PFOA is moderately accumulative in earthworms according to the soil conditions. Earlier, Zhao et al. (2012) reported that PFOA displayed a distinct bioaccumulation ability in E. fetida similar to that observed in the our study where PFOA with less fluorinated carbon atoms exhibited a distinguished accumulation pattern in all 3 soil types (Zhao, L et al. 2012). A sediment study on Chironomus plumosus explained that the same perfluoroalkyl chain length displayed varied BSAF kinetic value, i.e for PFOS and for PFOA (Xia et al. 2012). PFAS possess excellent water and oil repelling properties due to their failure to dissipate in octonal water systems (Conder et al. 2008; Liu, C et al. 2011). The commonly used octanol water partition coefficient K ow for lipophilic compounds is faulty in deciding bioaccumation of long chain PFAS. Similarly, the behaviour of PFAS differs from PAHs and PBDEs because normally hydrophilic compounds accumulate with log K ow 3-6. However, higher log K ow displayed a reduction in bioaccumulation (Sijm, Kraaij & Belfroid 2000). In PFAS the bioconcentration factors (BCFs) observed in carp increased with log K ow from PFOA (C = 8) to PFTA (C = 14). Also the BCFs of perfulorooctanoic acid (PFOA) and PFOS diversified by more than two orders of magnitude (PFOA BCF =<5.1 to 9.4; PFOS BCF = 720 to 1300) (Inoue et al. 2012). A previous study estimated that BSAFs from PFOA to PFDoA in E. fetida, using the exposure concentration of 100 ng g 1 were similar to those in the invertebrate organism Lumbriculus variegatus ( ) from PFOA to PFDoA (Lasier et al. 2011). More than a decade ago, (Weston, Penry & Gulmann 2000) introduced the idea that ingestion of contaminants can be the primary route of bioaccumulation of organic contaminants, along with adsorption and desorption procedure in the organism s gut. General PFASs are protein-loving compounds, and they show less interest in lipid contents (Jones, PD et al. 2003). Earthworms which comprised % of protein and % lipid were similar to Lumbriculus variegatus which obviously accumulated larger amounts of perfluorinated compounds as observed in our study. We can conclude that protein-rich organisms are subjected to higher accumulations of PFAS. 129

130 8.4.2 Effect on cell viability Any biological responses at the cellular level act as a biomarker for particular toxicants (Van Gestel, CAM & Van Brummelen 1996). Photometric assay, namely LDH cytotoxicty assay and MTT assay, were conducted in earthworms coloemocytes to evaluate the toxicity of PFOA. Our study concluded that the two assays were useful in assessing the cytotoxic potential of PFOA which affected cell vitality in all the treatments (Figs. 8.2 and 8.3). We reported that cell death induced by PFOA was prominent above 500 mg kg -1 in OECD soil and above 300 mg kg -1 in natural soils. The effects of PFOA were moderate compared to other PFCs. Previous invetigations have reported that PFOA seems low in toxicity to animals under controlled conditions but they were well-absorbed through oral and inhalation exposure, and to a lesser extent from dermal exposure (Kudo & Kawashima 2003). Adverse effects on plasma membrane released a stable enzyme, namely lactase dehydrogenase in all cell types. We used the release of LDH as a bioindicator for adverse effects on the cell membrane, quantified using the LDH cytotoxicity kit. Earlier studies reported that antibacterial effect in the coelomic fluid of Eisenia foetida was observed in using LDH release (Kauschke & Mohrig 1987). A detailed review on plasma membrane function indicated that most of the metazoan cells are prone to membrane disruption due mechanical stress. As a preventive measure they restructure themselves with endomembrane as its primary building block, and cytoskeletal and membrane fusion proteins as its catalysts (McNeil, Paul L. & Steinhardt 2003). Our study demonstrated the stress imposed on the organism which failed to stop the disruption because the release of LDH increased when the concentration also increased (Fig 8.2). The response may be due to the impacts on the dynamic cell or tissue level adaptation. Toxic sequel of oxidative stress at the subcellular level comprises lipid peroxidation and oxidative damage to DNA and proteins (Kelly et al. 1998). Tissue and cell level building blocks prevent forced bile seperation of plasma membrane either by guarding cells from damaging levels of stress, or, when this fails, they trigger the protein-based cables and linkages which pass through the plasma membrane as a safety measure. Thus oxidative stress generating the production of ROS in an uncontrolled manner blocks the antioxidant activity. It leads to important toxicological outcomes. PFOA is reported to induce oxidative stress and apoptosis in most animals like freshwater tilapia (Liu, Chunsheng et al. 2007a), planarian (Li, M-H 2008), and zebra fish (Liu, Yang et al. 2008). The MTT assay revealed that the laboratory PFOA exposed population showed moderate cellular damage compared to the control at cellular level (Fig 8.3). Therefore, acute 130

131 toxicity of PFOA exposed organisms was less significant at a smaller concentration, say below 500 mg kg -1, whereas a higher concentration induced significant cell impairments in neutral and alkaline soils. The OECD soil population indicated profound effects in cell vitality as shown in Fig. 3. We also observe a clear dose response relationship between the LDH and MTT assays. A similar pattern of cytotxicity was observed in Eisenia fetida on long-term exposure to cadmium, which was recorded using the MTT assay (Voua Otomo & Reinecke 2010). The current study reported the risk potential of PFOA is higher in OECD soil compared to natural soils. A study conducted on Eisenia andrei stated that the effect of cadmium on cell metabolism was stronger in OECD soil (Maleri et al. 2008). It is therefore illustrated that the photometric MTT and LDH assays can be used to evaluate the effects of PFOA on sub-cellular level in E. fetida Behavioral responses Effect of PFOA on locomotion Earthworms are highly sensitive organisms and they possess qualities to avoid stress induced by toxicants. In our study avoidance is visible even at lower concentrations whereas significant biological responses appear only at higher concentrations under acute toxicological testing conditions. Earthworms avoidance behaviour is exhibited in the form of altered locomotion pattern under contaminated stress at contaminated sites (Zheng, R & Li 2009). As the dose of PFOA increased there was much reduction in the crawling distance observed in Fig 8.3 (a-c), which was maintained under different soil conditions. PFOA exerted the similar level of stress in all the tested soils, since a significant effect was observed only above 200 mg kg -1 in all of them. Consequently the general pattern of inhibition in locomotion appeared to be the same. The locomotion process in earthworms is very sensitive to anthropogenic compounds because the produced significant behavioural changes at earlier concentrations where there was no significant effect on cellular responses (Figs. 8.2 and 8.3). Spontaneous behaviour (locomotion, rearing, and total activity), and habituation were noted in 2- and 4-month-old mice exposed to PFOS (0.75 or 11.3 mg), and PFOA (0.58 or 8.70 mg). The vulnerability of the cholinergic system was examined in a nicotine-induced spontaneous behaviour test in 4- month-old mice. Irregular unconstricted behaviour was observed in mice exposed to PFOS and PFOA, which is clear evidence for diminished or lack of habituation and hyperactivity in adult mice (Johansson, Niclas, Fredriksson, A & Eriksson, Per 2008). The mechanisms behind earlier reported behavioural defects could be because PFOA induced increased levels 131

132 of tau proteins in the hippocampus which is a protein responsible for proper brain functioning. It also modifies these proteins levels during a critical period of brain development (Johansson, Niclas, Eriksson, Per & Viberg, Henrik 2009). Thus the behavioural changes observed in the current study could be due to the effect of PFOA on the earthworms nervous system Effect on Cast production From the ecological perspective cast production is a promising bioindicator of xenobiotic stress. Our study reported that the effect of PFOA is similar all the soils which displayed clear dose dependent responses (Fig. 8.3 b, c). The most significant concentration in inhibiting cast production occurred above 300 mg kg -1 in neutral and alkaline soils. Redution in cast production is also a behavioural variation exhibited by earthworms due to the stress imposed by PFOA. Prenatal exposure to PFOA induced altered motor functioning in male and female offspring, Furthermore hyperactivity in males inside their home cage was recorded (Onishchenko et al. 2011). Cast prodcution in Lumbricus terrestris was very sensitive to pesticide toxcity. A significant reduction was noted at the normal application rate for pesticides (imidacloprid, carbaryl, methomyl) or without (ethyl-parathion and chlorpyrifos-ethyl). These also demonstrated a clear concentration effect response (Capowiez et al. 2010). Our behavioural test, based on estimations of CP, is that direct, rapid process that requires simple procedures and eradicates most of the issues observed when employing other behavioural standardised tests. More importantly, it is adjusted to earthworm species with higher ecological relevance (Lowe, CN & Butt 2005) and takes only 14 days. 14 day-lc 50 values for earthworms (Eisenia fetida) were estimated at the level of 1,000 mg kg -1 (dry weight) in the PFOA-exposed group (Joung et al. 2010). In our study a significant effect at concentration below LC 50 value was reported early. Thus PFOA significantly affect earthworms cast production Effect of PFOA on the wound healing capacity In this study, PFOA affected the wound healing capacity of test organisms in a dosedependent manner (Fig 8.4). On day 14 the highest concentrations causing significant effect on wound healing were 300 mg kg -1, observed on 7 th and 14 th day exposures, 200 mg kg -1 on 28 th day exposure from OECD soil PFOA treated group. In the neutral soil only 700 mg kg -1 of PFOA induced a significant effect but on the 14 th and 28 th days, exposures at a lower concentration of 300 mg kg -1 induced a significant effect. This trend was similar even in alkaline soil. From Fig 8. 4 we can conclude that duration of exposure plays a vital role in the 132

133 woud healing process. Consequently, wound healing is highly sensitive to exposure duration rather than concentration of PFOA. Wound healing in an organism is highly associated with the plasma membrane intergrity, and any distruption in it immediately kills the affected cell using a protective role. Yet animal cell plasma membranes are highly suspectible to mechanical shocks because they are not guarded by the cell wall. Also many tissue conditions impose heavy mechanical loads on the consistuent cells mainly at the physiological level (McNeil, P. L. & Steinhardt 1997). Our study reported that PFOA is active in inducing disruption to plasma membrane integrity; this was tested using LDH cytotoxicity (see Fig. 8.2). The main damage to the plasma membrane consisted of greatly compromising the wound healing process in the PFOA-treated population. Furthermore exposure duration is vital in promoting this action. Any mechanical damage to the plasma membrane elicits a Ca 2+ vesicle-vesicle fusion reaction that reseals the puncture by forming a new bilayer patch (Terasaki, Miyake & McNeil 1997). The mechanism of PFOA in influencing the wound healing process in earthworms is unclear but it may be speculated that they obstruct the Ca 2+ vesicle-vesicle fusion reaction. This is responsible for the rapid resealing that follows the breach in plasma membrane integrity. Apart from blocking the multifunctional calcium/calmodulin kinase, certain agents capable of affecting neutral tramsmission involving two synaptosomal-associated proteins - synaptobrevin and SNAP-2 - also inhibit the resealing process (Steinhardt, Bi & Alderton 1994). It can be concluded that kinetin and myosin are the motor proteins involved in exocytosis, i.e. successful sealing of the punctured cell membrane and in turn ensuring cell survival (Bi et al. 1997). Our study indcates that PFOA actively infects the plasma membrane but also inhibits proper resealing of the pathway by blocking exocytosis. Thus we can conclude that the PFOA affects the wound repair system of earthworms. They act as a candid biomarker for identifying the harmful effects of PFOA in a terrestrial ecosytem. 8.5 Conclusion PFOA accumulated in the earthworms body and caused many cellular damages. Ecological parameters of earthworms were affected due to acute exposure. 133

134 9 Application of biomarker battery for evaluation of Benzo(a)Pyrene effects on earthworms 9.1 Introduction The impacts of soil contamination are diverse in nature and are felt at the regional and international levels. These contaminants include mutagenic polycyclic aromatic hydrocarbons (PAHs) (Finlayson-Pitts & Pitts 1997). PAHs are known for their wide distribution mainly through incomplete coal combustion of organic matter because of biomass (e.g. wood and grass) burning and vehicle emissions in remote and urban locations (Yunker et al. 2002). Extensive studies have indicated that PAHs and their epoxides can cause a variety of toxic outcomes in lower organisms and higher organisms like mammals and humans. Removing these PAHs from the soil is very difficult as most physiochemical methods fail remediation process to the limitations involved (Samanta, Singh & Jain 2002). These global contaminants are even detected in the placental tissues and umbilical cord blood samples of mothers and new born babies, respectively. Their mutagenity has been proven by the identification of damaged DNA from the cord blood. These organic compounds need to be investigated in depth because they have serious consequences for people s health (Ravindra, Mittal & Van Grieken 2001). Most of the PAHs can induce tumours in human beings and smoking especially elevates PAHs levels (Mastrangelo, Fadda & Marzia 1996; Goldman et al. 2001). The distribution pattern of these compounds is interesting since they have more affinity towards fat composition in the body. Begin of neoplasm in fishes and other animals is mainly due to the activity of enzymes that are involved in the metabolism of PAHs. PAHs carcinogenicity is motivated by cytochrome P450 during the detoxification process. The main activity of P-450 is to displace the hydrocarbon at the same time encourage tumour promoters (Stegeman, John J & Lech 1991). Expression of CYPA1A has been identified in mesopelagic fishes from western north Atlantic sea (Stegeman, J. J. et al. 2001). From the above studies we can conclude that PAHs have the potential cause various ill 134

135 effects in all the biota in and around the contamination limits. Several studies have highlighted the carcinogenetic and mutagenic ability of these PAHs even in humans. For this reason the Unites States Environmental Protection Agency (USEPA) has included 16 PAHs on its priority list (Liu, K et al. 2001). The tumour- promoting quality of PAHs closely connected to the architecture of the compound. The number of fused benzene rings and also the capacity of the diol epoxide intermediates to bind the specific areas in DNA due to biotransformation of parent PAHs. Exposure limit to PAHs has been calculated by the WHO and in Air Quality Guidelines for Europe, wherein the unit risk is per ng/m3 of B[a]P. Guidelines on human health issue have reported that 0.1 ng m3-1 of B[a]P is enough to induce tumours in humans (Boström et al. 2002). The risk of phenanthrene and their derivatives causing cancer has been proved, and they are considered to be model compounds due their small structures. These are associated with bay region and K region, and it has been reported that ther carcinogenetic efficiency is less comparable to positive control benzo(a)pyrene 4,5-oxide (Bücker et al. 1979). Benzo(a)Pyrene poses a serious toxicological risk to both aquatic and terrestrial organisms. One study on mussels (Mytilus galloprovincialis) suggested that contamination with B(a)P leads to the formation of DNA adducts in the organism associated with malfunctioning of enzymes (Akcha et al. 2000). The principal mode of action of these PAHs having log K ow values lower than 5.5 or 6 is through altering the structure and function of cell membranes. This is termed as narcosis and it mainly occurs in terrestrial organisms (Jensen & Sverdrup 2003). Acute toxicity test for earthworms (OECD, 1984) were established to promote using earthworms in toxicological studies as key indicator species in environmental toxicology (Spurgeon, David J, Weeks & Van Gestel 2003). Biological responses from earthworms are reported to be a valuable biomarker for various PAHs such as pyrene and benzo(a)pyrene (Saint-Denis et al. 1999; Brown et al. 2004). Application of such biochemical responses as diagnostic and prognostic tools for assessing the toxic effect of contaminants has been reviewed in detail (Livingstone 1993). The toxicity tests suitable for soil organisms have been limited so far to acute toxicity and chronic bioassays (i.e. growth and reproduction). Thus, sublethal invertebrate tests are required for biomonitoring in the field context and for chemical testing under the controlled conditions (Saint-Denis et al. 2001). Thus our objective was to investigate the biological responses of the earthworm E. fetida exposed (B(a)P to different concentrations after 14 days. The aims were to firstly explain the effects at different biological organisation induced by B(a)P exposure and 135

136 secondly explore these responses as potential biomarkers for PAH-contaminated soil monitoring or for use in sublethal assays for chemical testing in the laboratory. In order to achieve our goal, earthworms were exposed to increasing concentrations B(a)P for 14 days employing the standard acute toxicity assay guideline (OECD, 1984). We investigated following markers (1) Mortality (2) Growth (3) Cellulase activity and accumulation (4) cytotoxicity and lysosome membrane stability and (5) behavioural changes. 9.2 Materials and Methods Refer chapter Results Acute toxicities highlighting survival, weight loss, cellulase activity and bioaccumulation B(a)P caused mortality in worms at 10 mg kg 1 and higher after seven and fourteen days of exposure. The earthworms resided at the rim of the container at the lowest B(a)P concentration (10 mg kg 1 ). The 14-d LC 50 was calculated for B(a)P in OECD soil, alkaline and neutral soil, which ranged from 73 to 150 mg kg -1. As the concentration of B(a)P in the soil increased, B(a)P level in the earthworms was elevated but the bioaccumulation factor decreased as the concentration of B(a)P increased, suggesting that the interaction between the earthworms and organic matter could further affect the biodegradation of B(a)P in the soil (Fig. 9.1a). The rate of accumulation in worms varied between the soils. It is apparent from the result that B(a)P accumulate more at lower concentrations 1-50 mg kg -1, because these concentrations showed higher BAF factor shown in the Fig (9.1c). The BAF values ranged from 0.3 to 0.9 in both the soil types. The growth inhibition by B(a)P of earthworm after 14 day (Fig 9.1b). Under the laboratory a condition, no weight loss was observed in control but weight gain was noted. The growth inhibition rate of earthworms exposed to B(a)P-treated soil was significantly higher than that of the controls in all the soils used. The decrease in weight was shown dose-dependent during 14-d exposure period. Weight loss was started at the lower concentration at 10 mg kg -1, whereas significant effect in weight reduction was appeared only above 50 mg kg -1. Visually we could predict that the health status of earthworms which looked appealing under control treatment while there was marked decline in the fitness of earthworms during the exposure period, as the concentration of respective B(a)P increased across each treatment group (mostly at higher concentration. This may indicate that the exposed earthworms were suffering from the stress imposed by B(a)P. The 136

137 data in the Figure 9.4 indicates the effect of different concentration of PAHs on the cellulase activity which was significant with that of control. The Dunnett t test (2 sided) showed statically significant difference in cellulase activity between control and all the treatments for 14 days (P<0.05). The cellulase activity was dose dependent in B(a)P exposed groups after14 days Fig (9.1c). The reduction in cellulase activity was significant even at 10 mg kg -1 which gradually decreased with increased level of B(a)P exposure in all the soil types, but higher concentration of B(a)P totally disrupted the cellulase activity which reflects that the organisms were starving for longer period. 137

138 Figure 9.1 Effect of B(a)P on earthworm (a) survival status, (b) rate of weight loss, (c) bioaccumulation of B(a)P and (d) cellulase activity estimated after 14 days of exposure in 3 soils (OECD, alkaline and neutral soils). Each point is the mean of three replicates. Error bars represent standard error (SE). 138

139 9.3.2 Cytotoxicity assay and membrane stability The effect of B(a)P on the cell vitality and viability (LDH and MTT assay) of earthworms is displayed in Fig. 9.2). Reduction in the cell viability was dose-dependently modified by B(a)P; cell death activity was initiated at 10 mg kg 1 followed by maximum apoptosis at high concentration treatment (300 mg kg 1 ) compared to the control group. Analysis by Dunnett 2 sided t test revealed that B(a)P concentration (P < 0.01) significantly affected the viability and cell vitality at 10 mg kg -1 in OECD soil and 50 mg kg -1 in neutral soil and alkaline soil. The effect of B(a)P on the membrane stability was shown in the Fig After 14 days the lysosome membrane stability was disturbed drastically with respect to increase in B(a)P concentration. Statistical analysis identified the significant effect on membrane stability above 50 mg kg -1 under both OECD and natural soils using Dunnet 2 sided t test (p<0.05). Figure 9.2 Cytotoxicity (%) observed in earthworm treated with B(a)P in alkaline, neutral and OECD soils. Cell vitality was determined as the mean of the absorption at 450 nm and expressed as % ± SD of the control. Reduction of the colorant MTT by coelomic cells after exposure for 14 days; cell viability was determined as the mean of the absorption at 570 nm and expressed as % ± SD of the control. 139

140 Figure 9.3 Lysosome membrane stability in earthworms exposed to of B(a)P measured after 14 days Lipid peroxidation, antioxidant capacity and DNA damage The effect of B(a)P on MDA content in earthworms is displayed in Fig After 14 d of exposure, MDA content in all treatment groups increased significantly compared to the control group. However, significant perturbations in MDA content were noted between control and B(a)P treated earthworms at 20 mg kg -1 in OECD sol and 50 mg kg -1 in natural soils. Statistical analysis revealed the concentrations (P < 0.01) had a significant effect on MDA content at exposer concentration above 20 mg kg -1. The effect of B(a)P on antioxidant capacity in earthworms is displayed in Fig Total antioxidant capacity in all treatment groups was inhibited by 1 14 d of phenanthrene exposure, and the inhibition was positively correlated with the increase in phenanthrene concentrations. The activity was increased at initial concentration up to 50 mg kg -1 followed by saturation at 100 mg kg -1 and declined at higher concentration later. Figure 9.5 shows the OTMs and % tail DNA of SCGE analysis of coelomocytes tested on the 14th day after treatment with different doses of B(a)P in different soils. The OTMs and % tail DNA at B(a)P dose ranging from 10 to 250 mg kg 1 were significantly higher than those of the controls (P < 0.05). DNA fragmentation, measured as OTM and % tail DNA, showed a dose response curve, therefore DNA damage increased as phenanthrene concentration increased. 140

141 Figure 9.4 Effect of B(a)P on earthworms total antioxidant capacity estimated as Each point is the mean of three replicates. Error bars represent standard error (SE) Tail DNA and OLive tail movement (%) % tail DNA - OECD OTM-OECD soil % tail DNA Neutral OTM-Neutral soil % tail DNA - Alkaline soil OTM-Alkaline soil Benzo(a)Pyrene concentration (mg/kg) Figure 9.5 DNA damage induced by B(a)P in earthworms measured using comet assay. Observed parameters were % tail DNA and olive tail movement. Each point is the mean of three replicates. Error bars represent standard error (SE) 141

142 MDA (nmol/mg) OECD soil Neutral soil Alkaline soil Figure 9.6 Benzo(a)Pyrene concentration (mg/kg) Effect of B(a)P on earthworm s lipid peroxidation estimated indirectly by measuring the MDA content produced inside the body. Each point is the mean of three replicates. Error bars represent standard error (SE) Behavioral response to B(a)P Results of the 5-d wound healing capacity are expressed as percent healing according to the concentration and duration of exposure (Fig 9.7c). The sensitivity of wound healing capacity is highly responsive to exposure concentration and duration. Both the concentration and exposure significantly affected the wound healing capacity of earthworms exposed in B(a)P in all the soils. Figures 9.7 and 9.8 clearly show the effect of B(a)P on the wound healing capacity of earthworms measured at different namely at 7 th day, 14 th day and 28 th day. B(a)P behaved in a different pattern with respect to different soil, here in OECD soil B(a)P was efficient in affecting the repairing mechanism after 7 th day, 14 th and 28 th days organism suffered seriously with the burden of B(a)P exposure to close the punctured wound during various exposure durations. Here significant effect was initiated at above 10 mg kg -1 identified using Dunnett 2 sided t test (P<0.01) under OECD soil. But in natural soil the effective concentration in inhibiting wound repair was 50 mg kg -1, after 28 th day the organism again failed to repair the wound so effects were severe as duration of exposure increases we even noticed death in worms. Locomotion and cast production was measured on day 14 th after acute exposure to B(a)P in earthworm. Distance covered by the earthworm in one minute along with the amount of cast they produced during experiment period is shown in Fig. 9.7a. Dunnette t test (2 sided) was carried out to identify the concentration inducing significant effect. B(a)P 142

143 induced significant effect even at lower concetration say around 10 mg kg -1 (0<0.05), there were significant effect even in cell vitality at this concetration (Fig. 9.7 a). Clear dose response curve was exihibited by earthworms exposed to B(a)P in affecting the locomotive skills of earthworms. Cast production was observed in natural soils, the amount of cast produced were weighed individually in all the treatments. The sieving was carried out immediately after 14 day exposure. PAH induced significant effect on cast production in the both neutral soil and alkaline soil (Fig 9.7a). Significant concentration reponse relationship was exihibited in organism exposed to B(a)P. The effect of B(a)P is stronger in inhibiting the organism from producing cast in treatments because even the lower dose 20 mg kg -1 (p<0.05) induced significant reduction in cast generation. So we can conclude that behavioural responses were more sensitive to B(a)P, because significant effects were appeared at even lower dose of B(a)P where there were no visible effects at cellular level of organisation in the earthworms exposed to natural soils. There was no mortality in the worms throughout the experiment period during earthworm avoidance behaviour test. When we exposed earthworm (Eisenia fetida) showed avoidance behaviour with respect to soil B(a)P concentration in all the soil types (Figure 9.7 b). B(a)P exhibited clear dose response relationship throughout the experiment with respect to their soil types displayed in the Fig. 9.7 b. B(a)P induced aversion to earthworms even at low concentration 10 mg kg -1 in OECD, alkaline and neutral soil. So it s clear that earthworms developed a well sensory system to detect B(a)P and avoid them as a survival mechanism. 143

144 Figure 9.7 (a) Mean (±SD) crawled distance by earthworm in one minute on 14 day and mean (±SD) of cast production (g/worm/day) on day 14 in earthworms exposed to B(a)P, (b) Avoidance behaviour of earthworms exposed to B(a)P (c) and (d) Percent of earthworms having complete healing after 5 d post wounding on 7 th, 14 th and 28 t -d exposure in 3 different soils. Sample size Is 20 for each group. 144

145 Wound healing in capacity (%) th day 14th day 28th day Benzo(a)Pyrene Concentration (mg/kg) Figure 9.8 Percent of earthworms having complete healing after 5 d post wounding on 7 th, 14 th and 28 th -d exposure to different concentrations of B(a)P in alkaline soil. Sample size is 20 for each group Reproduction effects Physical observation of test containers after 14 days showed significant negative effects in the reproduction in the treatment with B(a)P. B(a)P treated earthworms were active in producing cocoons only at 10 mg kg -1 in OECD soil and 20 mg kg -1 in natural soils, later there were reduction in reproduction rate. Trends in the juvenile emergence were also similar to the pattern of effect observed from the reduction in the coon production. The effect of B(a)P on the juvenile emergence appeared similar in OECD soil and natural soils, lower concentration (20 mg kg -1 ) were affective in inhibiting the juvenile emergence (Table 9.1). Although all worms were reproducing for the whole experiment period there was lethargic or dead adults predominantly at most concentrations after 14 days leading to a decline in cocoon production and juvenile emergence 145

146 Soil Type Concentration Cocoon production (28 days) Juvenile emergence (56 days) OECD 0 20± ± ±0.88* 12.33± ±0.58* 9.66±0.33* 50 10±0.58* 9±0.58* ±0.89* 6±0.52* 150 3±0.33* 2±1.20* ±0.56* 0.33±1.33* Neutral soil ± ± ± ±0.52* 13.66±0.58* 12.33±1.53* ±1.2*0 9.66±1.52* 100 9±1.0.58* 8±1.00* ±0.33* 5.66±1.15* ±0.34* 3±1.33* ±0.32* 1.33±0.56* Alkaline soil ± ± ± ±0.76* 20 17±0.58* 13.66±0.52* ±0.88* 9.66±0.53* ±1.20* 7.33±0.58* ±0.87* 5.66±1.53* ±0.67* 3±1.00* ±0.33* 2.66±0.33* ±1.36* 1±0.00* Table 9.1 Reproduction data of earthworms exposed to different concentration of B(a)P maintained in 3 soils. Parameters observed were Cocoon production on 28 th day and Juvenile emergence on 56 th day. 146

147 9.4 Discussion Survival, growth, cellulase activity and accumulation of B(a)P In our study, different biochemical factors along with physiological responses emerged as potential biomarkers for diagnosing the effects of B(a)P on E. fetida using three different soil conditions. The biochemical responses affected by the dose or the duration of exposure were mainly antioxidant capacity, lipid peroxidation, DNA and some physiological parameters like wound healing capacity. This analysis has shown that B(a)P was toxic to earthworms in both artificial and natural soils. Earthworms ingest the contaminant and accumulate them in their body either through the skin or digestive tract. Soil properties wield a large influence on the bioavailable fraction of contaminant to target organism. Some novel phenomena have been postulated regarding the fate of PAHs in terrestrial ecosystems, where bioaccumulation declined due to the variations in bioavailability associated with aging of PAH contaminated soil in enchytraeid (Ma, W, Immerzeel, J & Bodt, J 1995). Our results show that the B(a)P accumulated in the earthworms at higher rates when the concentration increased. Surprisingly the accumulations were low compared to other organic contaminants. For example the study conducted with 12 PAHs and 20 PCB contaminated soils reported that after 15 days exposure, the concentrations in earthworms did correlate with both the concentration and duration of exposure. The average biota-to-soil accumulation factor was times higher for PCBs compared to PAHs ( ) and were also independent of K ow (Krauss, Wilcke & Zech 2000). This correlated well with our results where the bioaccumulation factor (BAF) value from our study ranged between ( ). Accumulation of PAHs has been well investigated before in earthworms (Matscheko et al. 2002). Two earthworm species accumulated significantly different levels of PAH from the soil. Eisenia foetida and Lumbricus terrestris accumulated and μg g -1 total PAH, respectively, but neither of them accumulated measurable quantities (i.e. 5 or 6 ring PAHs). This suggests that not only is the accumulation of weathered PAHs species-specific; the bioavailability of individual PAH compounds is highly differentiated (Parrish et al. 2006). Theoretical data on PAH accumulation in worm tissues were available but they were too complex to compare due to several methods being used (worms species and age, incubation time, worm-to-soil ratio, contaminant concentration and age) and different data presentation values that may or may not be normalised to lipid content and the amount of organic carbon) (Parrish et al. 2006). Freshly spiked soil increased the bioavailability of PAHs reported in one 147

148 study that was conducted with earthworms of the Aporrectodea longa species. The authors suggested that uptake was more pronounced through their skin (Johnson, DL et al. 2002). The bioavailability of chemicals to earthworms can be altered strikingly by soil physico-chemical characteristics (Lanno et al. 2004). In order to develop a clear picture of modifying pictures, the measurement of chemical bioavailability to earthworms should be consistent and logical. Variations in the effect observed in our study may be due to changes in the availability of chemicals to earthworms. It is generally known that for a compound to be toxic in an organism, different factors play vital roles mainly in regard to exposure route, pore water concentration, and metabolism of parent compound in the body (Ma et al. 1998; Hoss et al. 2001). A previous study with PAHs reported that the sorption of organic pollutants onto the soil organic matter significantly influenced the bioavailability of the compound and toxicity (Weissenfels, Klewer & Langhoff 1992). In our study the difference in the biological response might be the effect of soil organic content in individual soil types. Organism s total uptake of chemicals in a terrestrial ecosystem depends on the amounts that were ingested rather than dermal contact. For example, PAHs such as B(a)P and phenanthrene accumulated through ingestion of PAHs contaminated sediment accounted more in total uptake of these compounds in oligochaete (Hyodrilus templetoni) (Lu, Reible & Fleeger 2004). However, one report on earthworms (Lumbricus rubellus) metal (Cd, Cu and Pb) uptake where the dermal intake plays vital role, indicated that the cuticles of earthworms were highly permeable to the chemicals (Vijver et al. 2003). The acute toxicity test in our study revealed that the 14-d LC 50 of B(a)P for E. fetida was mg kg 1 in natural soil. This LC 50 was higher than the 73 mg kg 1 in artificial soil, as determined by the same experiment. Surprisingly nil effect or only poor effects were observed in soil invertebrates within the concentration range tested (up to 947 mg kg 1 ) for B(a)P (Sverdrup, Line E. et al. 2007). Any prominent changes in weight can be a response due to chemical stress; they are highly linked to the energy budgeting of the body. The current study identified that weight gain in controls indicated that the soils in their original condition were good for promoting growth. The inhibition of growth results for B(a)P agreed with those documented for other organic pollutants. For example, inhibition of growth was sensitive in juveniles of Apporectodea caliginosa exposed to the organophosphorous insectides diazion and chlorpyrifos (Booth, LH & O'Halloran 2001). Weight loss was sensitive in Eisenia fetida earthworms exposed to deltamethrin and pyrene (Brown et al. 2004; Shi, Y et al. 2007). Weight gain was a very sensitive parameter in 148

149 determining the effect of B(a)P at the tested concentrations on E. fetida. B(a)P retarded earthworms growth which could be their survival tactic to overcome chemical stress. Earlier reports stated that the absence of earthworms in heavy metal contaminated sites was due to the effects of zinc on: firstly, the growth and maturation of juveniles; and secondly, the cocoon production rate of adult worms (Spurgeon, David J. & Hopkin 1996). These kinds of changes were also noticed in rainbow trout (Oncorhynchus mykiss) that lived in a metal contaminated site. They converted reduced food energy into biomass rather than reduced their food intake (Hansen, JA et al. 2004). When we compare this finding to our study we noted reduced food intake was indirectly caused by the reduction in cellulase activity. Cellulase is a digestive enzyme that can serve as a valuable biomarker for identifying chemical stress in earthworms (Luo et al. 1999; Shi, Y et al. 2007). Its presence in the earthworm s gut represents how they break down plant litter and other cellulosic materials. Cellulase plays a vital role in E. fetida, e.g. its activity was 7-fold higher than that in the gut of Metaphire guillelmi (Zhang, BG et al. 2000). This study shows that acute exposure to B(a)P reduced cellulase activity, indicating the disturbed condition in biochemical metabolism of earthworms. Reproductive toxicity will ultimately end in population decline of any organism, so it is a key factor in long-term ecological risk assessment. Additionally, carcinogens like PAHs will disrupt growth and development by promoting cancers causing death, so considering these factors when assessing the ecological risk monitoring is essential. A critical finding of this study was that B(a)P significantly reduced cocoon production and juvenile formation in earthworms. The reproductive toxicity of B(a)P has been previously reported in earthworms by identifying different expressions in the annetocin gene responsible for reproduction (Zheng, S et al. 2008). Ricketts et al. (2004) investigated annetocin expression level in earthworms exposed to heavy metal contaminated soil and reported that expression was 20 times less than the reference, and the cocoon production rate also fell 10- fold. These outcomes should be considered with regard to ecological risk and biotreatment process assessments. Reproduction endpoints (i.e. number of juveniles per hatched cocoon, survival of juveniles, inhibition of cocoon production, and hatchability) emerged as representative responses to estimated sub-lethal toxicity of energetic compounds compared to growth inhibition and survival of adult earthworms (Robidoux et al. 2000). 149

150 9.4.2 Biochemical responses and genetic damages In our study, different biochemical factors were assessed as potential biomarkers for diagnosing the effect of B(a)P on E. fetida. The biochemical responses affected by the dose or the duration of exposure were mainly the antioxidant enzymes, lipid peroxidation rates and DNA of earthworms. Information on time- and dose-dependent responses to B(a)P is required prior to using this organism to investigate the impact of PAHs. The concentrations of B(a)P considered in this study were environmentally relevant since these levels have been reported at roadsides and industrial sites. Also, high variability emerged in total PAHs ( PAHs) concentrations, ranging from less than 366 to 27,825 ng g 1 (Tang et al. 2005). Our study has shown that some responses were affected significantly by the treatment. Total antioxidant capacity and lipid peroxidation status were altered at the lowest dose and after 14 days exposure. However, the effect of dose of B(a)P and duration of exposure on the DNA were significant and appeared at 14 d of exposure which was indicated by the % tail DNA and OTM. It was also reported that Eisenia fetida possesses an effective detoxification system for PAHs, in that it guards against the toxic effects of these chemicals (Saint-Denis et al. 2001). It has been proved that that B(a)P can be metabolised either by enzymatic (cytochrome P450- dependent or not) or non-enzymatic pathways involving free radical reactions (such as lipid peroxidation) in earthworms. The excessive generation of free radicals was much scavenged by glutathione metabolism rather that GST because they appeared to be non-inducible in Eisenia fetida earthworms (Stokke & Stenersen 1993; Saint-Denis et al. 1999). Enzymatic defence against the ROS originated in the entire organism which includes SOD. This is essential in preventing lipid peroxidation by accelerating specific redox reaction and - converting oxy radicals into H 2 O 2 and O 2 and CAT. These elements are crucial in the transformation of H 2 O 2 into molecular water (Farber 1994; Hu, CW et al. 2010). Additionally B(a)P appeared to be less effective in triggering CAT activity in earthworms (Eisenia fetida) whereas significant inhibition in CAT activity was observed in the anterior and mid-region of earthworms exposed to olaquindox (Saint-Denis et al. 2001; Gao et al. 2007). These findings support our experiment where higher antioxidant activity in smaller concentrations occurred and declined as concentrations increased. Malondialdehyde (MDA) is an organic compound that is highly reactive and acts as a representative for oxidative stress produced from the reaction between free radicals and unsaturated fatty acids (Grundy & Storey 1998). They are responsible for the formation of inter and intra molecular protein links and causing cellular damages (Papadimitriou & 150

151 Loumbourdis 2002). Many investigations have suggested that the degree of lipid peroxidation was an essential factor in estimating the level of oxidative stress in organisms. In our studies, the MDA content in the treatment groups was higher than in the control group after 14 d of the experiment, indicating that the earthworms had sustained serious oxidative stress. These results suggest that earthworms cannot tolerate acute exposure to B(a)P and fail to combat oxidative damage. Oxidative stress could contribute to cancer in humans and animals through lethal processes such as membrane peroxidation, depletion of ions, protein separation, and DNA damage (Collins, Andrew & Harrington 2002; Mittler 2002). In the present study, the elevation in DNA injury was due to oxidative stress, indicating that ROS generation in tissues caused subsequent DNA breakages, or it was due to the activation of DNA repair mechanisms facilitated by B(a)P in earthworm coelomocytes. A number of studies have shown that ROS is the cause of DNA breakage, identified as nucleotides loss and many alterations in nucleotide bases (Cooke et al. 2003). Although cells have developed defence mechanisms to mitigate naturally induced DNA alterations, unlimited ROS generation can cause persistent injury to DNA. The source of the toxicological effect of B(a)P in the cell was their ability to serve as an oxidising agent once they entered the cells. They later reacted by reducing agent leading to myriad cellular damage like DNA adducts and DNA protein crosslinks (King et al. 1993) Membrane stability and cytotoxicity Assessment of integrity in the lysosome membrane through neutral red retention is regarded as a favourable biomarker in earthworms because it consists of simple analytical procedure and complex ecological realism. Xenobiotics consumed by earthworms first have to contend with the coelom, and injuries may appear early in the cells of the coelom. Lysosomal membrane stability can decline when stressful conditions return, and this is readily perceived in the NRR assay as a slow flow of the neutral red from the lysosomes into the surrounding cytoplasm (Sanchez-Hernandez, J 2006). Some researchers have determined that this biomarker is beneficial in anticipating the adverse effects on lifecycle traits (e.g., survival, growth, or reproduction). Week and Svendsen (1996) reported that NRR times in E. andrei exposed to Cu significantly declined when metal concentration in soil was 20 mg kg 1, whereas survival or changes in body weight were significantly inhibited at Cu concentrations as high as 320 mg kg -1 (Weeks, Jason M & Svendsen 1996). Similarly, our study reported a significant reduction in retention time at a lower concentration of 10 mg kg -1 of B(a)P. Soil 151

152 types do not have much influence on NRR assay, but the OECD soil appeared to be more reactive compared to natural soils. A study on earthworm (Apporrectodea caliginosa) demonstrated that NRR assay was sensitive to diazinon and chlorpyriphos after 28 days of exposure compared to survival and cocoon production (Booth, LH & O'Halloran 2001). Thus biomarkers such as NRR assay may be practical for calculating the internal consequence dose so that the bioactive contaminant fraction can be identified. This study compared the cytotoxicity of B(a)P and the results indicated that it caused severe cytotoxicity. Previously, sixteen polycyclic aromatic hydrocarbons (PAHs) were reported to cause stress to cell from the rainbow trout gill, RTgill-W1. Reports advised that exposure duration of 2 h or less was adequate for direct cytotoxicity, which appeared to be induced by general perplex of membranes (Schirmer et al. 1998). These authors also mentioned that 2 3 ringed compounds showed a higher cytotoxic effect whereas our study showed that benzo(a)pyrene induced more cell death in earthworms. Water solubility and lipophilicity are the 2 major properties regulating the direct cytotoxicity of PAHs and do so through the impact of PAH accumulation in earthworms membranes. Our results indicated that B(a)P accumulated more in earthworm tissues. Experimental evidence also showed that PAHs contained in the ambient airborne particulate matter act collaboratively in ROS production. These ROS are recognised as the main ingredient in cytotoxic and abnormal cell growth that leads to oxidative stress, pulmonary tissue injuries and DNA damage (Valavanidis, Fiotakis & Vlachogianni 2011). A possible cytotoxic effect of PAHs observed in our study could be the excessive generation of ROS. Various reports have highlighted the toxic effects with reference to PAHs. Most importantly, lipophilic hydrocarbons accumulate in the membrane lipid bilayer, and then infect the structural and functional characters of the cell membranes. Due to the accumulation of hydrocarbon compounds, the membrane suffers from membrane rupture, increases the flow of protons and ions in and out of the membranes. As a result of depletion of the proton intracellular ph homeostasis gets disturbed. Furthermore infection by lipophilic compounds on the lipid part of the membrane affects proteins embedded in the membrane (Sikkema, Jan, De Bont & Poolman 1995) Behavioral response Exposure to PAHs and environmental change can alter the behaviour of earthworms which was clearly demonstrated through our experiment by observing restricted locomotion, diminishing cast production, avoidance and inappropriate wound repair. One previous study indicated the route through which the behaviour might change in marine organisms is through 152

153 heightened metabolic burden, or info-disruption. Moving away from a contaminated environment was due to elevated CO 2 and reduced sea water ph (Briffa, de la Haye & Munday 2012). A recent study on the behaviour response of zebra fish (Danio rerio) reported that the PAH mixtures of different compositions can trigger disrupted behaviour in the organism resulting in endangered health (Vignet et al. 2014). Such kinds of behavioural abnormalities were visible during our experiment. The B(a)P was the most deleterious in infecting the ecological parameters. Vulnerability to single PAHs has been shown to degrade the swimming patterns of fish that coincided with our study. Specifically the B(a)P had a damaging effect on locomotion, cast production and the wound repair system (Gravato & Guilhermino 2009; Almeida, Gravato & Guilhermino 2012; Oliveira, Gravato & Guilhermino 2012). Similarly food intake and swimming efficiency significantly declined after exposure to BaP, with LOEC values 16 and 8 μ g L -1. Here B(a)P equal or higher than 8 μ g/l, caused detoxification capacity to shrink, an accumulation of liver BaP metabolites leading to lipid peroxidation, resulting in compromised growth and swimming capability of sea bass (Dicentrachus labrax) (Gravato & Guilhermino 2009). Similarly, in our study B(a)P was more effective even at a lower concentration of 10 mg kg -1 which disrupted behavioural responses in all soil types. Neurotoxic effects of BaP and pyrene have been identified earlier in rainbow trout and rockfish, suggesting that these compounds target the acetylcholine. This is an important neurotransmitter that suppresses the retinal cells neurite outgrowth, resulting in macroscopic changes in behaviour (Gesto et al. 2009; He, Chengyong et al. 2012). A similar cause could be responsible for varied behavioural traits identified in our study. Reconstruction of wound is an ancient process involving epidermal epithelial layers which are susceptible to external stress observed in C. elegans. Here surgical injury of the skin induced a large and sustained elevation in epidermal calcium. Epidermal calcium signals seem to specifically facilitate an actin-dependent mechanism of wound reconstruction (Xu, S, Hsiao & Chisholm 2012). Our study clearly showed that wound closure greatly compromised by PAHs load in earthworms. Deficiencies in the repair system might be due to in the generation of epithelial calcium being interrupted. The wound healing capacity in earthworms was sensitive to PCB because aroclor modifies earthworm coelomocytes and/or their interactions associated with antigen recognition and inflammation (Cooper, Edwin L. & Roch 1992). We could therefore speculate that PAHs also behave in a similar way as PCB in destroying the earthworms repair 153

154 9.5 Conclusion Overall, the results demonstrate that the battery of biomarkers utilised in this work could describe the evolution of a B(a)P induced stress syndrome in E. fetida. These biomarkers were able to reveal changes in the physiological status of the organism in response to sublethal concentrations of B(a)P. A comparison of growth effects and biomarker responses in this study suggests that the ill effects of BaP in earthworm manifest earlier at the suborganismal level than at individual level. Further, this study clearly showed that earthworm biomass was significantly affected even at low concentrations of BaP in the soils. 154

155 10 Effects of phenanthrene on the biological responses of earthworms under laboratory condition 10.1 Introduction The US Environmental Protection Agency monitors sixteen PAHs as priority pollutants, seven of which are reported as carcinogens. Of these certain PAHs are considered to be persistent organic pollutants (POPs) and persistent bioaccumulative chemicals (PBTs). Their classification also comprises polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins/dibenzo-p-furans (dioxins/furans), and several organochlorine pesticides such as DDT/DDE/DDD and chlordane (Ritter et al. 1995). Polycyclic aromatic hydrocarbons (PAH) and their alkyl homologs are ubiquitous in the environment and have been detected in sediments and soils worldwide. Studies have indicated that they stubbornly remain in all the examined locations, and in terms of quantity PAHs abundance increased near urban centres. It can be concluded that anthropogenic combustions of fossil fuels and other substances are major sources of contamination (Laflamme & Hites 1978; Ribes et al. 2003). The number of rings plays a major role in their environmental stability because of their remarkable binding with lipids and their stable chemical structure facilitates easy adsorption into soil organic matter (Means 1980). Dispersion of PAHs in soils is highly influenced by runoff and dust generation, so contamination in soil is considered to be one of the important gateways for pollution in air and sediments (Mai et al. 2003). Polycyclic aromatic hydrocarbons (PAH) are more prevalent in soil, yet information on ecotoxicological risks imposed by these compounds to life in soil is scanty. Academic and regulatory perspectives on PAHs focus more on human health-related issues but currently ecotoxicological aspects of these PAH gradually have garnered much attention among the scientific community (Douben 2003). A recent investigation on expression profiles of PAHs in rat liver identified characteristic molecular signatures for each PAH, and reported 1183 genes as a candid molecular signature recognising PAHs and its genotoxicity (Jung et al. 155

156 2013). The molecular evidence suggests that exposure to environmental carcinogens such as polycyclic aromatic hydrocarbons and aromatic amines heightened the risk of causing cancer in human beings (Perera 1997). A previous investigation into the toxicity of PAHs reported that PAHs with log K ow values 5.2 (i.e., naphthalene, acenaphthene, acenaphthylene, anthracene, phenanthrene, fluorene, pyrene, and fluoranthene) significantly reduced the survival and reproduction of springtail Folsomia fimetaria (Sverdrup, Line E., Nielsen & Krogh 2002) and acted as immune toxicants in Mytilus edulis (Coles, Farley & Pipe 1994). Some of the ecological parameters such as reproduction were sensitive to PAHs exposure in terrestrial isopods namely Oniscus asellua and Porcellio scaber (Van Brummelen, Van Gestel & Verweij 1996). PAHs accumulate in the fat tissues. For example, an in situ study on 19 polycyclic aromatic hydrocarbons (PAH 19 ), focusing on their distribution, biotransformation and flux in the food chain in seston-blue mussel (Mytilus edulis) and common eider duck (Somateria mollissiima ) reported that they were well distributed in the gallbladder, liver, adipose tissue and egg of the duck (Broman et al. 1990). This has led to more research into their chemical and biological properties in the environment, and the process through which these physiological consequences occurred. The importance of earthworms as a key indicator species has been extensively studied previously so they are regarded as a vital part of soil fauna in the terrestrial ecosystem (Edwards 2004b). They are employed in environmental risk assessments in order to identify the bioavailability and toxicity of both organics, mainly PAHs (Bergknut et al. 2007; Gomez-Eyles, Collins & Hodson 2010) and inorganics, mainly heavy metals (Hobbelen, Koolhaas & Van Gestel 2006; Svendsen, C. et al. 2007) because their route of intake includes both direct skin contact and also by ingestion of the contaminated food, soil and pore water (Lanno et al. 2004). PAHs are highly bioaccumulative in earthworms tissue apart from this they also induce mortality in the earthworm community (Jonker et al. 2007). Toxicological and biochemical responses of earthworm Lumbricus rubellus to pyrene have been well documented (Brown et al. 2004). The lethal effects of PAHs on earthworms in contaminated soil have been calculated by biological responses such as avoidance, survival, growth, reproduction, and protein content. The responses of earthworms to sub-chronic exposure have also been well documented. These responses are considered as biomarkers which play a vital role in the ecotoxicological risk posed by pollutants. Indeed they provide early warning signals with respect to pollutant exposure (Shi, Y et al. 2007). Extensive studies concluded that PAHs can 156

157 produce higher level of reactive oxygen species (ROS) in earthworms. Reactive oxygen species comprises of superoxide anion radical O - 2, hydrogen peroxide (H 2 O 2 ), and the highly reactive hydroxyl radical ( OH). Earthworms are naturally built with an efficient defence system that scavenges the ROS. The main role of antioxidant enzymes includes transformation of O 2 into H 2 O 2 by SOD, while CAT decomposes H 2 O 2 into oxygen and water. Elevated level of ROS not only induces an oxidative stress but also damages the lipid peroxidation in cells of earthworms. Being an important product of lipid peroxidation cellular malondialdehyde (MDA) content can act as a biomarker for level of lipid peroxidation (Pan, Ren & Liu 2006; Li, XY et al. 2010). Therefore, the alterations in the biological response after the contact with pollutants (such as anti-oxidant enzymes and lipid peroxide levels) in the earthworm were due to the oxidative stress in the whole organism caused by the metabolism of the pollutants. In this study, the effect of phenanthrene and its bioaccumulation in earthworms as influenced by the soil properties was investigated under laboratory conditions. Acute and sub-chronic toxicity tests were performed to evaluate the effects of phenanthrene on mortality, biomass, anti-oxidant enzyme activities, lipid peroxidation, reproduction, cytotoxicity and behavioural responses in the mature earthworms. The aim of this investigation was to generate more information to better understand the complex effects of phenanthrene on earthworms in the soil ecosystem Material and Methods Refer chapter Results 157

158 Figure 10.1 Effect of phenanthrene on earthworm (a) survival status, (b) rate of weight loss, (c) bioaccumulation and (d) cellulase activity estimated after 14 days of exposure in OECD and neutral soils. Each point is the mean of three replicates. Error bars represent standard error (SE). 158

159 Acute toxicities highlighting survival, weight loss, cellulase activity and bioaccumulation Phenanthrene caused mortality in worms at concentrations ranging from10 mg kg 1 to 600 mg kg -1 after seven and fourteen days of exposure. The earthworms resided at the rim of the container above 100 mg kg 1 phenanthrene. The 14-d LC 50 for phenanthrene was calculated in OECD soil and neutral soil, which ranged from 220 to 280 mg kg -1. As the initial concentration of phenanthrene in the soil increases, phenanthrene content in the earthworms was also increased but the bioaccumulation factor decreased as the concentration of phenanthrene increased. This suggests that the interaction between the earthworms and organic matter could further affect the biodegradation of phenanthrene in the soil (Figure 10.1a). It is also important to state the rate of accumulation varied between the OECD soil and neutral soil. PAHs seem to accumulate more at lower concentrations from 1-50 mg kg -1, because they showed a higher BAF value (Figure 10.1c). The BAF value ranged from 0.02 to 0.5 in both the soil types. The effect of phenanthrene on earthworm biomass after 14 d exposure is shown in (Figure10.1b). Under the laboratory conditions, there was no weight loss in the control worms, in fact, their weights increased during the 14-d exposure period. The growth inhibition rate of earthworms exposed to phenanthrene-treated soil was significantly higher compared to controls in both the soils used. The decrease in weight was shown to be dose-dependent over the 14-d exposure period. Weight loss was started even at the lower concentration at 10 mg kg -1, whereas significant effect in weight reduction was appeared only above 150 mg kg -1. Visually we could predict the health status of earthworms by appearance since there was marked decline in the fitness of earthworms during the exposure period, as the concentration of respective phenanthrene increased across each treatment group (mostly at higher concentration). This may indicate that the exposed earthworms were suffering from the stress imposed by phenanthrene. The data in the Figure 4 indicates the effect of different concentration of PAHs on the cellulase activity which was significant with that of control. The Dunnett t test showed statically significant differences in cellulase activity between control and all the treatments for 14 days (P<0.05). The cellulase activity was dose dependent in phenanthrene exposed groups after14 days (Figure10.1c). The reduction in cellulase activity was initiated as early at 10 mg kg -1 but no sharp decline was noticed up to 500 mg kg -1, but higher concentration of phenanthrene totally disrupted the cellulase activity suggesting that the organisms were starving for longer period. 159

160 Figure 10.2 (a) Cytotoxicity (%) observed in earthworms treated with phenanthrene exposed in 2 soils. Cell vitality was determined as the mean of the absorption at 450 nm and expressed as %± SD of the control. Reduction of the colorant MTT by coelomic cells of earthworms after exposure for 14 days; cell viability was determined as the mean of the absorption at 570 nm and expressed as % ± SD of the control. Figure 10.3 Lysosome membrane stability in earthworms exposed to phenanthrene measured after 14 days Cytotoxicity assay and membrane stability The effect of phenanthrene on the cell vitality and viability (LDH and MTT assay) of earthworms is displayed in Fig Reduction in the cell viability was dose-dependently 160

161 modified by phenanthrene; cell death was low at low concentration treatment (10 mg kg 1 ) followed by maximum cell death at high concentration treatment (500 mg kg 1 ) in neutral soil and 350 mgkg -1 in OECD soil compared to the control group. Analysis by Dunnett s t test revealed that phenanthrene concentration (P < 0.01) significantly affected the viability and cell vitality at 50 mg kg -1 in OECD soil and 150 mg kg-1 in neutral soil. The effect of phenanthrene on the lysosome membrane stability was shown in the Fig After 14 days the membrane stability was declined drastically in response to increase in phenanthrene concentration. Statistical analysis identified the significant effect (p<0.05) on membrane stability above 150 mg kg -1 and 250 mg kg -1 under both the soils. Figure 10.4 Effect of phenanthrene on earthworm total antioxidant capacity. Each point is the mean of three replicates. Error bars represent standard error (SE) 161

162 % tail DNA - OECD OTM-OECD soil Tail DNA and Olive tail movement (%) % tail DNA Neutral OTM-Neutral soil Phenanthrene concetration (mg/kg) Figure 10.5 DNA damaged induced by phenanthrene in earthworms measured using comet assay. The parameters evaluated were % tail DNA and olive tail movement. Each point is the mean of three replicates. Error bars represent standard error (SE) 4.5 MDA content (nmol/kg) OECD soil Neutral soil Phenanthrene concetration (mg/kg) Figure 10.6 Effect of phenanthrene on earthworm lipid peroxidation estimated indirectly by measuring the MDA content produced inside the body. Each point is the mean of three replicates. Error bars represent standard error (SE) Lipid peroxidation, antioxidant capacity and DNA damage The effect of phenanthrene on lipid peroxidation in earthworms is shown in Fig After 14 d of exposure, MDA content in all treatment groups increased significantly compared to the control group. However, significant perturbations in MDA content were 162

163 noted between control and phenanthrene-treated earthworms at 100 mg kg -1 in OECD soil and at 200 mg kg -1 in neutral soil. Statistical analysis revealed the concentrations (P < 0.01) had a significant effect on lipid peroxidation. The effect of phenanthrene on antioxidant capacity in earthworms is displayed in Fig Total antioxidant capacity in all treatment groups was inhibited by 14 d of phenanthrene exposure, and the inhibition was not positively correlated with the increase in phenanthrene concentrations. The activity was increased at initial concentration up to 50 mg kg -1 followed by saturation at 150 mg kg -1 and declined at higher concentration of phenanthrene. Fig 10.5 shows the OTMs and % tail DNA of SCGE analysis of coelomocytes tested on the 14th day after treatment with different doses of phenanthrene in 2 soils. The OTMs and % tail DNA at phenanthrene dose ranging from 50 to 700 mg kg 1 were significantly higher than those of the controls (P < 0.05) compared to OECD and neutral soil. DNA fragmentation, measured as OTM and % tail DNA, showed a dose response relationship thus DNA damage increased with an increase in phenanthrene concentration. 163

164 Figure 10.7 (a) Mean (±SD) crawled distance by earthworm in one minute on 14 day and mean (±SD) of cast production (g/worm/day) on day 14 in earthworms exposed to phenanthrene (b). Avoidance behaviour of earthworms exposed to phenanthrene (c) and (d) Percent of wound healing 5 d post wounding after 7, 14 and 28-d exposure in neutral and OECD soils. Sample size is 20 for each group. 164

165 Behavioral response Results of the 5-d wound healing capacity are expressed as percent healing according to the concentration and duration of exposure as shown in the Fig 10.7c. The sensitivity of wound healing capacity is highly responsive to exposure concentration and duration. Both the concentration and exposure significantly affected the wound healing capacity of earthworms exposed to phenanthrene in both soils. Figures 10.7c andd clearly show the effect of PAHs on the wound healing capacity of earthworms measured at 7 th day, 14 th day and 28 th day. The fate and behaviour of phenanthrene varies depending on soil properties and concentrations. In this study, OECD soil was more efficient in wound healing after the 7 th day but after day 14 worms managed to recover although the effect was still observed at higher concentration, after 28 th day exposure organism well equipped for the stress and recovered well compared to 7 th and 14 th day exposure. Yet significant effect was initiated at above 50 mg kg -1 as indicated by Dunnett 2 sided t test (p<0.01) for OECD soil exposure. But in neutral soil 7 th day exposure was severe in affecting the repairing process with worms failed to recover at higher concentration, after 14 days worms recovered well compared to effects observed at 7 th day exposure, but after 28 th day the worms again failed to repair the wound so effects were severe as duration of exposure increased. Here significant effect (p<0.05) in wound healing was noted above 150 mg kg -1. Locomotion and cast production were measured on day 14 th after acute exposure to phenanthrene in earthworms. Distance covered by the earthworm in one minute along with the amount of cast they produced during experiment period is shown in Fig 10.7a in OECD and neutral soils. Phenanthrene induced significant effect (0<0.05) even at lower concetration of about 50 mg kg -1, although there was no significant effect in cell vitality at this concetration (Fig 10.7a). With regards to locomotive skills, a clear dose response curve was exihibited by earthworms exposed to phenanthrene. Cast production was observed in neutral soil, the amounts of cast produced were weighed individually in all the treatments. The sieving was carried out immediately after 14 day exposure. PAHs induced significant effect on cast production in neutral soil. A significant concentration - reponse relationship was exihibited in worms exposed to phenanthrene. The effect of phenanthrene was stronger in inhibiting the worm from producing cast because even the lower dose 50 mg kg -1 (p<0.05) induced significant reduction in cast generation. The results show that behavioural responses were more sensitive to phenanthrne, because significant effects appeared at lower dose where there were no visible negative effects at cellular level of organisation in the earthworms. 165

166 There was no mortality in worms throughout the experiment period. However, earthworms showed avoidance behaviour with respect to soil phenanthrene concentration in both the soils (Figure 10.6). Phenanthrene exhibited clear dose response relationship throughout the experiment with respect to their soils (Fig 10.7). As shown in the Fig. 10.7b phenanthrene showed attraction at initial concentration (10 mg kg -1 ) but significant effect was noted at 150 mg kg -1 in OECD and neutral soils (P<0.05). Avoidance test was less sensitive for phenanthrene. 166

167 Soil Type Concentration Cocoon production (28 days) Juvenile emergence (56 days) OECD ± ± ± ± ± ±1.00* ±1.00* 12±0.89* ±2.64* 10.67±0.67* ±1.22* 8.33±0.34* ±1.15* 6.33±2.30* ±0.58* 3.66±0.58* Neutral soil ± ± ± ± ± ±1.00* ± ±0.89* ±1.00* 8.67±1.73* ±0.57* 6±1.52* ±0.89* 3.33±1.15* ±1.20* 2.67±0.58* ±1.27* 1.33±0.67* Table 10.1 Reproduction of earthworms exposed to different concentration of phenanthrene maintained in 2 different soils namely OECD and neutral soils. Parameters observed were cocoon production on 28 th day and juvenile emergence on 56 th day. 167

168 Reproduction effects Physical observation of test containers after 14 days showed significant reduction in reproduction as the phenanthrene concentration increased. Earthworms were less active in producing cocoons at higher concentration of phenanthrene around 150 mg kg -1 in OECD soil and 250 mg kg -1 in neutral soil. Trends in juvenile emergence were also similar to the pattern observed earlier from the reduction in cocoon production. The effect of phenanthrene on the juvenile emergence appeared to be more in OECD soil compared to natural soil (Table 10.1). Although all worms at lower PAH concentrations were alive, active, and reproducing for the whole experimental period, there were lethargic or dead adult worms at higher concentrations after 14 days. This finally led to a decline in cocoon production and juvenile emergence Discussion Survival, weight loss, cellulase, bioaccumulation and reproduction This study has shown that phenanthrene was toxic to earthworms in both artificial and natural soils. Earthworms come into contact with the pollutants more frequently, ingesting them that account for accumulation of toxicant in the body tissue facilitated through their skin and digestive tract. This has been which well documented in in the studies with phenanthrene and fluoranthene (Ma, W, Immerzeel, J & Bodt, J 1995). Also bioaccumulation declines in the organism in aged soils due to lower bioavailability. Importantly, earthworms produce strong binding with PAHs under situation of food restrictions (Ma, WC, Immerzeel, J & Bodt, J 1995). This agrees with our study since we conducted acute toxicity assay without food which might explain higher accumulation of phenanthrene in earthworms. Also the two soils (natural and artificial) induced differential toxicities and accumulation pattern in worms which might be due to the differences in the bioavailable fraction of phenanthrene in 2 soils. The average biota-to-soil accumulation factor (BSAF) of PAHs in earthworm ranges from 0.13 to 0.41 in contaminated field study which is in close agreement with the data obtained in the current study (Krauss, Wilcke & Zech 2000). This finding may vary slightly due to leaching and runoff and they were considered to be important factors facilitating contaminant disappearance which is not applicable in our study. The amount of organic matter also plays vital role in degradation of PAHs. According to the present experimental data, bioaccumulation of phenanthrene in earthworms was greater at initial soil concentrations leading to higher BAF values. Similar findings were reported in other studies with earthworms and PAHs (Krauss, Wilcke & Zech 2000). The acute toxicity test in this 168

169 experiment showed that the 14-d LC 50 of phenanthrene for E. fetida was 220 mg kg -1 in OECD soil and 280 mg kg -1 in natural soil. This is in contrast to previous reports which reported that LC 50 was mg kg 1 in natural soil and mg kg 1 in artificial soil (Zhao, Z 2007; Wu et al. 2011). The changes might be due to the variation in exposure and bioavailability of phenanthrene to earthworms. It was reported that the PAHs with low K ow values 5.2 (i.e., naphthalene, acenaphthene, acenaohthylene, anthracene, phenanthrene, fluorene, pyrene and fluoranthene) exhibited more pronounced toxicity towards survival and reproduction of soil dwelling springtail, Folsomia fimetaria (Sverdrup, Line E., Nielsen & Krogh 2002). They also postulated that the toxicity order increases with the increasing lipophilicity of the substance. Even the study on toxicity of 8 polyaromatic compounds to red clover (Trifolium pratense), ryegrass (Lolium perenne) and mustard (Sinapsis alba) reported that the EC 50 value for phenanthrene and pyrene were mg kg -1 and mg kg -1, respectively. The effect on seedling growth was more sensitive than on seedling emergence (Sverdrup, Line E et al. 2003). A similar finding was noted in the current study where the effect on growth is more sensitive than the survival of the organism. Changes in the biomass can be candid biomarkers for the induced toxicant stress which helps to understand the link between the chemical s impact and energy dynamics in the body ultimately; this denoted as growth retardation. Weight gain observed from the control worm proved that the nutrient status in the soil was good enough for normal growth of earthworms. Weight loss data for phenanthrene were similar to other findings with respect to organic contaminants. Sub lethal concentration of dieldrin induced significant reduction in the weight loss in a dose dependent manner in earthworms E. fetida (Reinecke, A & Venter 1985). Low molecular weight PAH like pyrene was also efficient in inducing significant variation in biomass observed in Lumbricus terrestricus (Brown et al. 2004). Growth was very responsive in estimating the impact of phenanthrene at the tested concentrations on E. fetida. Phenanthrene retarded the growth, which could indicate how they survive under stress (i.e., restriction in ingestion to avoid the toxins). This mode of survival was implemented by earthworms in many cases to avoid other contaminants like heavy metals. Even pesticides like endosulfan, lindane and deltamethrin strongly affected the body of isopod Porcellio dilatatus and earthworms mainly through modifying the feeding scenario responsible for growth inhibition (Ribeiro et al. 2001; Shi, Y et al. 2007). The observed changes in the cellulase activity in the current study clearly support this. Cellulase is an active digestive enzyme in earthworms which facilitates the digestion of cellulose material in the food they ingest as reported in earthworms like Pheretima hilgendorfi, Pontoscolex corethurus, Hormogaster elisae (Zhang, B et al. 1993; 169

170 Garvıń et al. 2000; Nozaki et al. 2009). Most of the toxicological reports on pesticides suggested that cellulase activity act as a valuable biomarker of chemical stress in earthworms on exposure.imidachloprid and RH-5849 were reported to be inhibitory to cellulase activity in E.foetida after exposure period (Luo et al. 1999). These findings support our findings considering even the lower concentration of phenanthrene inhibited the activity of cellulase which was reflected as reduced growth rate in earthworms after exposure period. Life history pattern such as reproduction are highly crucial in regulating the population growth in an ecosystem (Kammenga et al. 2003), these sub-fatal outcomes have likely to curtail the health status of earthworm populations in PAHs contaminated soil, leading to deterioration of population dynamics. The toxicities of sediment-associated phenanthrene, fluoranthene and diesel fuel in estuarine copepod Schizopera knabeni were investigated and reported that they were detrimental to reproduction coupled with reduction in fitness of the organism (Lotufo 1997). Also Eisenia fetida maintained in artificial soil spiked with phenanthrene (Phe), pyrene (Pyr), fluoranthene (Flu), or benzo(a)pyrene (Bap) reported that expression of annetocin and TCTP genes were affected at 10.0 mg kg 1 phenanthrene, also witnessed weight loss and cocoon production of the worms where no significant differences on cocoon production were reported (Zheng, S et al. 2008). Ricketts revealed that reduction in expression of annetocin level in earthworms exposed to heavy metal contaminated soil was associated with reduction in the cocoon production (Ricketts, HJ et al. 2004a). Presently in our experiment phenanthrene appeared to be more effective in disturbing the reproductive system of earthworms Lysosome membrane stability, cell cytotoxicity Assessment of the lysosome membrane integrity through neutral red retention is regarded as a favourable biomarker in earthworms because of its simple analytical procedure and complex ecological realism. Xenobiotics consumed by earthworms first have to contend with the coelom, and injuries may appear early in the cells of the coelom, Consequently, lysosomal membrane stability may be disturbed due to stress, and this can be readily perceived in the NRR assay as a slow flow of the neutral red from the lysosomes into the surrounding cytoplasm (Sanchez-Hernandez, J 2006). Some researchers have determined that this biomarker is beneficial in that it anticipates adverse effects on lifecycle traits (e.g. survival, growth, or reproduction). A study by (Weeks, Jason M & Svendsen 1996) reported that NRR in E. andrei exposed to Cu significantly declined when metal concentration in soil was 20 mg kg 1, whereas survival or changes in body weight were significantly inhibited at 170

171 Cu concentrations as high as 320 mg kg -1. In our study lysosome membrane was vulnerable to phenanthrene shock because it was much disrupted even at lower concentrations of phenanthrene. Thus NRR assay could serve as a practical tool for calculating the internal consequence dose which ideally manifests the bioactive contaminant fraction. In previous studies, exposure to PAHs damaged the lysosome membrane structure and vitality in mussels (Lowe, D, Soverchia & Moore 1995; PROGRESS 1995). PAHs from indoor dust have been reported to induce potential cytotoxicity in human cell lines (Kang, Cheung & Wong 2010). It has also been proved that 16 PAHs have a cytotoxic effect on a cell line from the rainbow trout gill. Of these, 2 to 3 ringed PAHs were found to be directly cytotoxic (Schirmer et al. 1998). Many cyclic hydrocarbons, for example aromatics, cycloalkanes, and terpenes, are noxious to microorganisms by targeting the cytoplasmic membrane so that toxic action can occur (Sikkema, J., De Bont & Poolman 1994). In our study, the disturbance of membranes appears to be the reason for observed cytotoxicity and the maintenance of membrane stability was estimated using 2 cytoxicity assays. Water solubility and lipophilicity of PAHs also influence cell death. Of the 16 PAHs those with high water solubility have been reported to be cytotoxic (Mackay, Shiu & Ma 1997). Different PAHs could react in either an additive or non-additive pattern to promote cell death. The toxicity of several aromatic hydrocarbons was due to the combination of different compounds interfering with membrane permeability as observed in Daphnia magna (Deneer et al. 1988; Munoz & Tarazona 1993) and inhibitory reaction in mitochondrial respiration (Beach, AC & Harmon 1992) Antioxidant capacity, lipid peroxidation and DNA damage Alterations in organisms biological processes to chemical stress are considered to be signals of pollution in the environment (Wilson & Crouch 1987; Song, Y. et al. 2009). Biochemical responses like enzymatic activities have been regarded as early warning signs of environmental pollution. The main role of these enzymes was to act as a protective shield against fatal events caused by excessive generation of reactive oxygen species (ROS) in an organism. Viable cells convert oxygen into water through their electron transport chains and induce antioxidant enzymes like SOD and CAT to scavenge the ROS produced inside the body as a reaction to the metabolism of toxicants (Farber 1994; Hu, CW et al. 2010). The results of our study showed that worm total antioxidant capacity (TAC) was induced initially and declined later when the concentration of phenanthrene rose. The data for TAC suggested that the earthworm suffered more burdens from the phenanthrene at higher concentrations. 171

172 A similar effect has been observed in the study conducted on earthworms exposed to different concentrations of phenanthrene, where the SOD and CAT activities were considerably affected at higher concentration exposure (Wu et al. 2011). Long-term exposure to phenanthrene failed to produce significant differences in CAT activity, possibly due to a longer exposure time leading to an adaptive mechanism triggered by the higher amount of phenanthrene. This has been confirmed in heavy metal exposure studies (Posthuma & Van Straalen 1993). The reaction between free radicals and unsaturated fatty acids in cellular membranes yields a product called malondialdehyde. The excessive free radicals combine with the free amino acids of protein to form inter and intra protein cross-links which lead to fatal events in cells. Thus the production of MDA can be reflect cellular damage caused by the free radicals (Papadimitriou & Loumbourdis 2002). This acts as a sensitive indicator of stress in some higher organisms like toads, Scaphiopus couchii (Grundy & Storey 1998). Numerous studies suggested that the amount of MDA was an essential factor in determining the amount of oxidative stress in organisms. In our experiment, the MDA content in the treatment groups was higher than in the control group after 14 days of the experiment, suggesting that the earthworms had survived toxic oxidative stress. Acute exposure to phenanthrene induced significant differences in MDA content among the treatment groups at higher concentrations. These results indicate that earthworms suffered from serious oxidative damage. Malondialdehyde (MDA), one of the most common carbonyl products of lipid peroxidation, reacts with DNA to form adducts to deoxyguanosine, deoxyadenosine, and deoxycytidine (Marnett 1999). Oxidative stress leads to a plethora of sickness vents such as membrane peroxidation, ion loss, protein breakdown, and DNA injury, which might cause tumours in humans or animals (Collins, Andrew & Harrington 2002; Mittler 2002). Herbicide atrazine caused oxidative stress and DNA damage in earthworms which appeared to be an important mechanism of its toxicity to earthworms (Song, Y et al. 2009). These findings agree with our study in which phenanthrene induced oxidative stress and in turn this led to the production of MDA which finally affected the DNA of earthworms. Earlier findings have proved that ROS is the origin of DNA impairment caused by strand breakage, loss of nucleotides, and various alterations of the nucleotide bases (Cooke et al. 2003) Behavioral responses Exposure to PAHs and environmental change can alter the behaviour of earthworms which is clearly demonstrated in the current experiment, in terms of observed effects on 172

173 locomotion restriction, diminishing cast production and inappropriate wound repair. An earlier report indicated the routes through which the behaviour might be changed in marine organisms when they experienced heightened metabolic burden, info-disruption and avoidance behaviour. They moved away from the contaminated environment of elevated CO 2 and reduced sea water ph (Briffa, de la Haye & Munday 2012). The PAH mixtures of different compositions can trigger disruption in the behaviour of the organism resulting in degraded health status as reported in zebra fish (Vignet et al. 2014). Our results indicated that earthworms can detect phenanthrene and avoid them by implementing various survival modes such as restricted movement, reduction in cast production and impaired wound healing. Vulnerability to single PAHs has been shown to disturb the swimming pattern in fish, and similarly in our study phenanthrene had a deleterious effect on locomotion, cast production and wound repair system (Gravato & Guilhermino 2009; Almeida, Gravato & Guilhermino 2012; Oliveira, Gravato & Guilhermino 2012). Three PAH compounds fluorene (FE), phenanthrene (PHE), and pyrene (PY) induced changes in their locomotory activities and social behaviours in fish (Gonçalves et al. 2008). The observed behavioural alterations in earthworms are mostly accepted as a toxicity response to organic narcotic compounds, which is also called narcosis (Hsieh, Tsai & Chen 2006). The main reason for this phenomenon is due to impairment in the cell membrane because the toxicant attacks the lipid content in them (Van Wezel & Opperhuizen 1995) but this process is rectifiable. The responses vary depending on the type of compound and exposure causing either complete or partial narcotisation resulting in symptoms such as neural imbalance, immune-system retardation, and anatomical malfunctions (Heintz, Short & Rice 1999; Vines et al. 2000). Reconstruction of the wound is an ancient process involving epidermal epithelial layers which are susceptible to external stress as observed in C. elegans. Here surgical injury of the skin induced a large and sustained elevation in epidermal calcium. Epidermal calcium signals seem to specifically facilitate an actin-dependent mechanism of wound reconstruction (Xu, S, Hsiao & Chisholm 2012). Our study clearly shows that wound healing was less extensively compensated due to phenanthrene load in earthworms. Cast production in earthworms has been identified as a new behavioural biomarker for toxicity testing (Capowiez et al. 2010). Avoidance behaviour acts as an early warning signal for PAHs contamination assessment where estuarine amphipod (Eohaustorius estuaries) avoided PAHs contaminated sediment at a concentration where no mortality was noted (Kravitz et al. 1999). Our results 173

174 correlated well with the findings that phenanthrene at lower concentrations induced avoidance in earthworms. The avoidance test is a behavioural test with several merits (easy, rapid and cost effective) but its main disadvantage is that it cannot estimate toxicity. Instead it calculates the amount of repellence as identified in the experiment on earthworms exposed to imidacloprid (Capowiez, Y. & Bérard, A. 2006) and thus is regarded as a measure of habitat modification. Improvements in this method have been proposed and implemented to better define the link between avoidance and toxicity (Sanchez-Hernandez, J 2006). Our behavioural test, based on the measure of CP, is direct, quick, and does not require special instruments. It is able to and overcome most of the disadvantages listed above that compromise other standardised behavioural tests. Furthermore it has been designed for earthworm species with higher ecological relevance (Lowe, CN & Butt 2005). Our results show that earthworms were moderately sensitive to phenanthrene in terms of their behavioural responses Conclusion This study shows that the phenanthrene induced toxicological and behavioural responses in earthworm such as survival, weight loss, cellulase, bioaccumulation, reproduction effects, antioxidant capacity, lipid peroxidation, DNA damage, lysosome membrane stability and cell cytotoxicity can serve as valuable biomarkers for phenathrene toxicity. 174

175 11 Differential responses of biomarkers in the body of earthworms exposed to pyrene under laboratory condition 11.1 Introduction Polycyclic aromatic hydrocarbons (PAHs) are pervasive contaminants produced by human activities in which combustion sources are responsible for over 90% of their environmental existence. Non-combustion processes such as the production and use of creosote and coal-tar, although poorly estimated, are possible primary and secondary sources (Howsam & Jones 1998). In total sixteen PAHs were termed significant pollutants by the US Environmental Protection Agency and importantly, seven of them have been reported to cause cancer. PAHs can be ranked as persistent organic pollutants (POPs) and persistent bioaccumulative chemicals (PBTs), consisting of polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins/dibenzo-p-furans (dioxins/furans), and a handful of organochlorine pesticides such as DDT/DDE/DDD and chlordane (Ritter et al. 1995). The pestilential nature of PAHs has been reviewed in detail, revealing an important link between the occurrence of PAHs and cancer. High occupational risk to PAHs exposure in relation to aluminium production, coal gasification, coke production, iron and steel foundries, tar distillation, shale oil extraction, wood impregnation, roofing, road paving, carbon black production, carbon electrode production, chimney sweeping, and calcium carbide production has been reviewed (Boffetta, Jourenkova & Gustavsson 1997). PAHs contain diverse molecular structures with 2 or more fused rings whose physicochemical properties such as water solubility varies with respect to the configuration of molecules, from 31 mg/l for naphthalene (2 rings) and 290 ng/l for benzo(g,h,i)perylene (6 rings) (Sims & Overcash 1983). The availability of PAHs and Polychlorinated Biphenyls (PCBs) to earthworms (Lumbricus terrestris ) has been reported with respect to the urban environment (Krauss, Wilcke & Zech 175

176 2000). Generally in eukaryotes the removal of xenobiotics is executed through two phases. Phase I employs the cytochrome P450 enzyme system to introduce a functional group, namely hydroxyl or sulphonyl, to non-polar compounds. On the other hand phase I reaction can start the development of metabolites that are more lethal than the parent compounds. Glutathione S-transferase (GST) which is a phase II enzyme responsible for physiological removal of compounds, is involved in the introduction of large polar, water-soluble element to the metabolites of phase I metabolism; it facilitates detoxification. Four families of cytochrome P450 enzymes (1, 2, 3 and 4) are motivated after contacting any lipophilic xenobiotic compounds and are essential for detoxifying xenobiotic compounds in annelids (Roos, PH et al. 1996; Solé et al. 1996; Lee, R. F. 1998; Van der Oost, Beyer & Vermeulen 2003). Radical cations speculated to be the positively charged species accountable for covalent bonding with chemical species of cellular macromolecules such as proteins, donate electrons for pairing (Cavalieri & Calvin 1971; Cavalieri & Auerbach 1974). An earlier study supported the view that the lipid-rich microsomal membranes are possible targets of injury in cells exposed to active oxygen species (Afanas'ev et al. 1993). Oxidative injuries to cells and tissues were obstructed by antioxidant enzymes such as catalase and others of small molecular weight (such as glutathione). Catalase (CAT, EC ) reduces the amount of hydrogen-peroxide. Glutathione is essential in the detoxification of xenobiotics in a given sequence: (1) acts as a substrate for glutathione peroxidase (GPX, EC ) enzymes; (2) it helps in the reformation of ascorbate and tocopherol; (3) it may be joined to electrophilic xenobiotics by glutathione transferase; and (4) possibly associated with pro-oxidants such as transition metals or sulphydril groups (Flohe 1982; Christie & Costa 1984; Meister 1988; Hasspieler, Behar & Digiulio 1994). Earthworms (E. fetida andrei) were provided with a well-developed defence system involving glutathione, glutathione-related enzymes and catalase to deal with PAHs (Saint- Denis et al. 1998). The SCGE assay using earthworms (Eisenia fetida) coelomocytes has recently emerged as an attractive biomarker for estimating exposure to genotoxic compounds (Di Marzio et al. 2005). A study on genotoxicity of PAHs on earthworms suggests that PAHs were responsible for DNA damage and poor repair process in earthworms treated with soil irrigated by waste waters (Qiao et al. 2007). For contaminants targeting the DNA of an organism, estimating the mutation designated as number of DNA strand breaks (DSBs) is an encouraging option. The comet assay has potential in identifying DSBs in individual 176

177 eukaryotic cells after in vivo or in vitro treatment and is recommended as an effective biomarker for evaluating the major modification in the raw material of an evolving gene (Faust et al. 2004). The repercussions of treatment and breakdown of mutagens in an organism can be well determined by observing either the DNA adduct formation or number of DNBs (Farmer et al. 2003; Husgafvel-Pursiainen 2004). Other notable or favourable conditions responsible for the formation of DSBs are high alkaline condition which enables inadequately functioning to recoup DNA adducts, crosslinks and alkali-labile sites. One promising approach is to determine mutagenicity as an estimation of % tail DNA compared to tail length and other parameters which are susceptible to variation in electrophoresis condition (Pfuhler & Wolf 1996; De Boeck et al. 2000). The aim of this study was to evaluate the potential of specific earthworm biomarkers to indicate soil toxicity as a supplement to the DNA damage caused by the PAHs Benzo(a)Pyrene, phenanthrene and pyrene in earthworms (Eisenia fetida) after acute exposure for 14 days in three different soil conditions. Earthworm biomarkers in terms of physiological response, biochemical response, and genetic biomarkers used were: lipid peroxidation, total antioxidant capacity and DNA damage. Thus our objective was to investigate the biological responses of the earthworm E. fetida exposed to pyrene at different concentrations after 14 days. The aims were: firstly, to explain the effects at different biological organisation induced by pyrene exposure; and secondly, to explore the potential for applying these responses as biomarkers for PAHcontaminated soil monitoring or for use in sub-lethal assays for chemical testing in the laboratory. This means identifying the ecological parameters. In order to achieve our goal, earthworms were exposed to increasing concentrations of pyrene for 14 days using the standard acute toxicity assay guideline (OECD, 1984). We investigated the following markers: (1) mortality (2) growth (3) cellulase activity and accumulation (4) cytotoxicity and lysosome membrane stability and (5) behavioural changes Materials and methods Refer chapter 3 177

178 11.3 Results Accumulation, survival, growth and cellulase activity Survival of worms in the contact tests is presented in Fig. 11.1a. All worms survived in the controls (indicating there the solvent pre-treatment had no effect) and also at the lowest pyrene concentrations. At higher pyrene concentrations there was a steady concentration related to decrease in survival. The LC 50 for the soil test ranged between mg kg -1 with respect to soil types. The wide difference between soils was observed in the experiment where OECD soil exhibited the most toxicity. Mortality in Eisenia fetida occurred after 14 days exposure and importantly higher concentration of pyrene induced notable death in earthworms. However, a dose dependent decrease in survival was clearly observed in all the treatments displayed in Figure Pyrene also had a significant effect on earthworm weight change (P < 0.05). Post-hoc comparison indicated a significant difference between treatments and control, with greater weight loss at 1200 mg kg -1 when compared to 50 mg kg -1. Lower concentration of pyrene (10-50 mg kg -1 ) increased the weight similar to that of control in earthworms. It is important to note that the rate of accumulation varied between the soils. Results show that PAHs accumulate more at lower concentrations from 1-10 mg kg -1, and they showed higher BAF factors (Fig. 11.1c). The BAF values ranged from 0.05 to 0.5 in all the soils. Visually the earthworms looked healthy using the control treatment but there was a marked decline in the fitness of earthworms exposed to pyrene as its concentration increased. This indicates that the exposed earthworms suffered from the stress imposed by pyrene mostly at higher concentrations. Figure 11.1c illustrates the effect of different concentrations of pyrene on cellulase activity, which declined significantly from 50 mg kg -1 to 1200 mg kg

179 Figure 11.1 Effect of pyrene on earthworm (a) survival status, (b) rate of weight loss, (c) bioaccumulation of phenanthrene and (d) cellulase activity estimated after 14 days of exposure in three soils OECD, alkaline and neutral soils. Each point is the mean of three replicates. Error bars represent standard error (SE). 179

180 Cytotoxicity assay and membrane stability The effect of pyrene on cell vitality and viability (LDH and MTT assay) of earthworms is displayed in Fig Reduction in cell viability was dose-dependently modified by pyrene; cytotoxicity was low at a low concentration treatment (200 mg kg 1 ) followed by maximum toxicity at a high concentration treatment (2500 mg kg 1 ) compared to the control group. Analysis by Dunnett 2-sided t test revealed that pyrene (P < 0.01) significantly affected the viability and cell vitality at 200 mg kg -1 in OECD soil, while it was 300 mg kg -1 in neutral and alkaline soils. The effect of pyrene on membrane stability is depicted in Fig After 14 days membrane stability declined gradually when the pyrene concentrations increased. Statistical analysis identified a significant effect on membrane stability above 300 mg kg -1 in all the soils using Dunnett 2-sided t test (p<0.05). (%) Vitality to that of control Cell vitality in OECD soil Cell vitality in neutral soil Cell vitality in alkaline soil LDH cytotoxicty in OECD soil LDH cytotoxicty in Neutral soil LDH cytotoxicity in Alkaline soil Pyrene Concentration (mg/kg) LDH cytotoxicity (%) Figure 11.2 (a) Cytotoxicity % observed in earthworm treated with pyrene in three soils. Cell vitality was determined as the mean of the absorption at 450 nm and expressed as %± SD of the control. Reduction of the colorant MTT by coelomic cells of earthworms after exposure for 14 days; cell viability was determined as the mean of the absorption at 570 nm and expressed as %± SD of the control. 180

181 Neutral red retension time (min) OECD Soil Neutral Soil Alkaline Soil Pyrene Concentration (mg/kg) Figure 11.3 Lysosome membrane stability in earthworms exposed to pyrene measured after 14 days Lipid peroxidation, antioxidant capacity and DNA damage The effect of pyrene on MDA content in earthworms is displayed in Fig After 14 d of exposure, MDA content in treatment groups increased significantly compared to the control group. However, significant perturbations in MDA content were noted between control and pyrene treated earthworms at 300 mg kg -1 in OECD and natural soils. Statistical analysis revealed these had a significant effect on MDA content at an exposed concentration above 300 mg kg -1. The effect of pyrene on antioxidant capacity in earthworms is indicated in Fig Total antioxidant capacity in treatment groups was inhibited by day 14 of phenanthrene exposure especially at a higher concentration, and this inhibition was positively correlated to the increase in pyrene concentrations. The activity increased exponentially up to 500 mg kg -1 and reached its maximum at 1000 mg kg -1. Thereafter it remained more or less constant up to 2500 mg kg -1. Figure 11.5 shows the OTMs and % tail DNA of SCGE analysis of coelomocytes tested on the 14 th day after treatment with varied doses of pyrene in different soils. The OTMs and % tail DNA at pyrene dose ranging from mg kg 1 were significantly higher than those of the controls (P < 0.05). DNA fragmentation, measured as OTM and % tail DNA, showed a dose response curve with DNA damage increasing as pyrene concentration increased. 181

182 0.40 mm trolox equivalent OECD soil Neutral soil Alkaline soil Pyrene concentration (mg/kg) Figure 11.4 Effect of pyrene on earthworm total antioxidant capacity in 3 soils. Each point is the mean of three replicates. Error bars represent standard error (SE) Tail DNA and Olive tail movement (%) % tail DNA - OECD OTM-OECD soil % tail DNA Neutral OTM-Neutral soil % tail DNA - Alkaline soil OTM-Alkaline soil Pyrene concetration (mg/kg) Figure 11.5 DNA damage induced by pyrene in earthworms measured using comet assay. The parameters evaluated were % tail DNA and olive tail movement 182

183 MDA (nmol/mg) OECD soil Pyrene concetration (mg/kg) Figure 11.6 Effect of pyrene on earthworm lipid peroxidation estimated indirectly measured as the MDA content produced inside the body. Each point is the mean of three replicates. Error bars represent standard error (SE) Behavioural response to phenanthrene Results of the 5-d wound healing capacity are expressed as percent healing according to the concentration and duration of exposure (Fig. 11.7c). The sensitivity of wound healing capacity is highly responsive to exposure concentration and duration. Both the concentration and exposure significantly affected the wound healing capacity of earthworms exposed to pyrene in different soils. Figures 11.7 and 11.8 clearly show the impact of pyrene on the wound healing capacity of earthworms measured at different times, i.e. 7 th day, 14 th day and 28 th day. Pyrene behaved in a different pattern with respect to the nature of soils. In the OECD soil pyrene did affect the repair mechanism after the 7 th day but after day 14 organisms managed to recover. A larger effect due to a higher concentration was noted after the 28 th day of exposure but the organisms were well equipped for stress and recovered well compared to the 7 th and 14 th day exposures. Yet a significant effect was observed at above 200 mg kg -1 when using the Dunnett 2-sided t test (P<0.01). However, in natural soil the most effective concentration in inhibiting wound repair was 300 mg kg -1 after 28 th day, when the organism again failed to repair the wound. Thus the effects worsened when duration of exposure increased. Locomotion and cast production were measured on day 14 after acute exposure of earthworms to pyrene. Distance covered by the earthworm in one minute along with the amount of cast produced during the experiment is shown in Fig 11.7a in OECD, alkaline and neutral soils. Dunnette t test (2-sided) helped to identify the concentration that induced the 183

184 significant effect; pyrene did so only above 200 mg kg -1 (0<0.05). A clear dose response curve was exhibited by earthworms exposed to pyrene that affected the locomotive skills of earthworms. Cast production was observed in natural soils and the amounts of cast produced were weighed individually in all treatments. Sieving was carried out immediately after 14 days exposure. Pyrene induced a significant effect on cast production in the both neutral soil and alkaline soil as shown in Fig 11.7a. Significant concentration reponse relationship was demonstrated in organisms exposed to pyrene. Dunnett t test (2-sided) helped to identify the concentration that induced the significant effect. The effect of pyrene is weak in preventing the organism from producing cast because a significant effect was only observed above 300 mg kg -1 (p<0.05). We can conclude that behavioural responses were more moderately sensitive to pyrene. There was no mortality in the worms during the avoidance behaviour test. Earthworms showed avoidance behaviour with respect to soil pyrene concentration in all the soils (Fig 11.7 b). Pyrene attracted earthworms at a lower concentration up to 50 mg kg -1 in the OECD, alkaline and neutral soils (Fig 11.7b). It is therefore, clear that earthworms developed a good sensory system that detected pyrene and preferred it at lower concentrations and avoided it at higher concentrations. 184

185 Figure 11.7 (a) Mean (±SD) crawled distance by earthworm in one minute on 14 day and mean (±SD) of cast production (g/worm/day) on day 14 in earthworms exposed to pyrene (b). Avoidance behaviour of earthworms exposed to phenanthrene (c) and (d) Percent of earthworms Eisenia fetida having complete healing after 5 d post wounding after 7 th, 14 th and 28 th -d exposure in 3 soils. Sample size is 20 for each group. 185