Bioaccumulation and depuration in sea urchins Paracentrotus lividus
(Lebanon) and Heliocidaris erythrogramma (Australia)
Waste deposit mountain
Carol M.S. Sukhn
MSc
Subtidal Ecology and Ecotoxicology Evolution and Ecology Research Center School of Biological, Earth and Environmental Sciences University of New South Wales Sydney Australia
Thesis submitted in fulfilment of the requirements for the degree of Doctor of Philosophy at the University of New South Wales
School of Biological, Earth and Environmental Sciences (BEES) Faculty of Science March 2013
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THE UNIVERSITY OF NEW SOUTH WALES
Thesis/Dissertation Sheet
Surname or Family name: Sukhn
First name: Carol Other given name/s: N/A
Abbreviations for degree as given in the University Calendar: PhD
School: Biological, Earth & Environmental Sciences Faculty: Science
Title: Bioaccumulation and depuration in sea urchins Paracentrotus lividus (Lebanon) and Heliocidaris erythrogramma (Australia).
Abstract
With toxicants being introduced daily into the environment, monitoring becomes an essential tool for the protection of human health and biota. The sea urchin, Paracentrotus lividus, was surveyed off the Lebanese coast to assess whether proximity to a major source of pollution affected abundance, biomass, size and metal content of the urchin. Metal concentrations in four body parts of the urchin did not vary with distance from the pollution source, nor did size, abundance and biomass, which were more directly affected by an unplanned closure to the fishery in 2007. The distribution of the urchin is therefore not considered a sensitive bioindicator of contamination, however, the contaminant loads measured are considered a direct concern with regards to human health, as consumption of urchin roe is currently unregulated.
To better understand the bioaccumulation of contaminants in the body parts of P. lividus, laboratory exposures to organics (PAHs, OCPs), inorganic metals (Ni,V, Cd, Pb) and bacteria were undertaken. The experiments also determined the bioconcentration factor in body parts of urchins, where the roe was found to accumulate the most contaminants out of all body parts (test, spines and teeth). The urchins accumulated PAHs, toxic metals and OCPs. P. lividus was also tested for its ability to depurate toxicants either naturally (field translocation) or in the laboratory. Sea urchins eliminated Ni, V, Pb, dieldrin and 4,4‘DDE at different rates while the depuration of Cd was not observed. Field depuration was more rapid and consistent than laboratory depuration and could be used by the aquaculture industry to meet food safety regulations.
The Australian sea urchin Heliocidaris erythrogramma was tested for its ability to bioaccumulate metals and PAHs in a laboratory exposure study. Gender differences in accumulation were observed and an initial baseline of metals was established for the organism at one site in Sydney Harbour. The levels of metals in the roe were within the food safety regulations except for lead. P. lividus and H. erythrogramma are good bioaccumulators and their roe may be used as potential biomonitors for short periods of time after oil spills or pulsed exposure events.
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Acknowledgments
Acknowledgment is about giving thanks to people, events and things, which helped in the completion of this thesis. So....
My first thank you goes of course to my supervisor Dr Emma Johnston for accepting my application to the doctorate, for guiding me throughout the thesis and for keeping me ―logical‖ all the time. Also for being so courageous to visit the sampling sites in Lebanon at a time of turmoil just to make sure that everything was set up correctly. I also thank you for being a statistical wiz and for teaching me lots and lots of statistics; I am forever indebted to you. Many thanks go to my co-supervisor Dr Imad Saoud for his good guidance, his ―occasional‖ patience and for setting up a good marine laboratory allowing for these studies to happen. Oh! And a special thank you for paying for most of the analyses.
I thank Dr Nadim Cortas for establishing the Environment Core Laboratory and for his valuable comments and review of the thesis. I also thank Dr Ghazi Zaatari for granting me some time off to write this thesis.
The fieldwork would never have happened without the help of the NGO ―Cedars for Care‖ president Mrs Iffat Chatila and her volunteers. A special thank you to the best skipper Mr Ahmad Iskandarani (Cedars for Care) for the many hours of rolling out ropes in the temporal variation experiment over the past seven years. These experiments would not have happened without the priceless help of the Lebanese Army and its divers. To every single one of you (pilots, navy and seals) I give my thanks and my appreciation. I know the diving conditions were so hard at some sites over the years, but your thoroughness in the implementation of instructions and your punctuality was very much appreciated.
To the staff of the Environment Core Laboratory at the American University of Beirut (AUB): Sanaa Fayad, Osanna Nashalian, Rawia Masri and Lina Beydoun, thank you for your help and patience. Rosaline Sislian: can I ever thank you enough? I don‘t think so. Without you I would not have made it on time. To Dr Youssef Mneimneh, thank you for allowing me to use your laboratory whenever I needed it. To the volunteers who helped me every now and then with field or laboratory work, a thank you from the heart.
To the staff of the marine biology laboratory at AUB, Joly Ghanawi especially, thank you for your help throughout all the experiments particularly for the many times you were dragged in on Saturdays and Sundays to help change water and clean tanks.
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To Dr Jane Williamson and her team at the Marine Laboratory of Macquarie University, my sincere gratitude for allowing me the use of your premises to run the bioaccumulation experiment on the Heliocidaris erythrogramma.
To my family: Saidé, Milad, Joseph, Marina, Sophie, Gabrielle, Elena, Suzie, Dany, Mary-Jo, Charbel Ray, Lu-Jane, Julie, Seamuss, Jack, Hala, Jihane, Nick, Jamal, John, Anna, Rayan, Nada, Sol, Raoul, Mary, Siham, thank you for the support and love. You are my rock. Cousin Nada, even though you were expecting me to go to London to write up this thesis just so I could mention you in my acknowledgments (as you eloquently joked about) well here you go: I ended up writing my thesis down under, but you are still thanked for the years of genuine love and great support. To Tony, Olga and Joseph Bou Ghosn Sukhn, Saideh Bou Ghosn Sukhn Abdel Ahad, Julia, Mamie and Youssef Maroun, Joseph Shakour and Marc Hlavacek: you are always on my mind.
Dr Nada Sabra, Dr Akram Ghantous, Dr Marc Hlavacek, thank you for your friendship and encouragement. Thank you Dr Jihane Sokhn Ball (sis) for all your comments and caring while reviewing my chapters. Thank you Dr David Roberts, Dr Katelyn Edge, Dr Katherine Dafforn, Dr Ceiwen Pease, Dr Jane Williamson and Dr Luke Hedge for reviewing some of the chapters.
To my extended family down under (Becharas and Marouns): Elie, Mary, Laura, Alexandra, Natalie, Nabiha, Joseph, Latifah, Assad, Charbel, Julie, Joe, John, Nabiha, Tony, May, Daniela, Julian and Luc. Thank you from the heart for your hospitality. I couldn‘t have done it without your support. Elie and Mary Bechara: Thank you for everything.
Marine scientists often face dilemmas in pinning down which of two probable causes (pollution or overfishing) is behind an organisms‘ near extinction, though not many would have the chance to completely eliminate a probable cause. I was faced with this same dilemma: What was the cause of the extremely low abundance of sea urchins in Sidon? Then in the aftermath of the war and oil spill in July 2006, a very stringent sea blockade was imposed on the Lebanese south coast until January 2007. This blockade was so efficient that no fishing whatsoever occurred in the area during that time. This eliminated one of the probable causes for low abundance of the sea urchins and allowed me to draw some conclusions. So I guess fate should also be acknowledged!
Finally a word to the sea urchins: Paracentrotus lividus: Thank you for being a great subject of study. I wish you a speedy recovery in our neck of the wood. I hope to stumble on dozens of you per square meter next time I go sampling (like it used to be when I was a kid). I wish you a biomass of more than 1000 g wet weight g-1. I promise I will never complain anymore of your presence on our rocky sea floor (and for stinging my feet). Now I know it is better to have some of you then to have none at all. H erythrogramma: you are yummy! I hope yours would be a better fate.
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Table of Contents Acknowledgments ...... v Table of Contents ...... vii 1.General Introduction ...... 1 Aims of study ...... 4 Thesis outline ...... 6 2.Temporal variation in abundance, biomass, size and metal content of the sea urchin Paracentrotus lividus in the South of Lebanon ...... 7 Abstract ...... 7 Introduction ...... 8 Material and Methods: ...... 10 Survey Methods and Design ...... 10 Tissue contamination measurements...... 13 Statistical Methods ...... 14 Comparison with published reports ...... 15 Results ...... 16 Density, biomass and size ...... 16 Sidon 2005-2009 ...... 16 Sidon-Tripoli Comparison 2007 ...... 18 Metal bioaccumulation in body parts ...... 19 Sidon 2007 ...... 19 Sidon 2007 -2008 ...... 22 Sidon and Tripoli 2007 ...... 24 Background metal levels in seawater and percentage body burden ...... 25 Metals ...... 25 Chemical and Bacteriological ...... 28 Discussion ...... 29 Density, Biomass and Size ...... 32 Metals in Tissues ...... 36 Conclusion ...... 41 3.Bioaccumulation and Bioconcentration factors for a cocktail of inorganic and organic toxicants in the sea urchin Paracentrotus lividus ...... 43 Abstract ...... 43 Introduction ...... 44 Bioaccumulation & Bioconcentration factor ...... 44 Common marine contaminants ...... 44 The study organism ...... 47
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Aims ...... 48 Materials and Methods ...... 48 Collection of Specimen (Sea Urchins) ...... 48 Experimental Set-up and Spiking (Exposure) ...... 49 Trial 1 ...... 49 Trial 2 ...... 53 Sampling and Analyses ...... 53 Metal analysis for trials 1 and 2 ...... 54 Quality Control for metals ...... 54 Organic methods OCP-PAH ...... 54 Organic Quality control...... 57 Microbiological methods ...... 58 The Bioconcentration Factor ...... 58 Data treatment and statistical analyses ...... 59 Results ...... 59 First trial ...... 59 Size, Total Weight and Roe Weight ...... 59 Lead and Cadmium in tissues ...... 61 Dieldrin and DDE in roe ...... 65 Acenatphthene and Pyrene in roe ...... 66 Second trial...... 67 Size, Total Weight and Roe Weight ...... 67 Lead, Cadmium, Nickel and Vanadium in tissues ...... 69 Anthracene, pyrene and phenanthrene in tissues ...... 71 Bioconcentration factor and metal concentration ratio in sea urchins body parts ...... 72 Discussion ...... 77 Metals ...... 77 Organics ...... 80 Compounds retrieved from the Gas Chromatography Mass spectrometry ... 84 Conclusion: ...... 87 4.Depuration of Cd, Pb, Ni, V , dieldrin and 4,4’DDE from the sea urchin Paracentrotus lividus in the laboratory and in the field...... 88 Abstract ...... 88 Introduction ...... 89 Aims ...... 91 Method ...... 92
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Collection of Sea Urchins ...... 92 Experimental set-up and Spiking (Exposure) ...... 92 Laboratory Depuration ...... 96 Translocation depuration ...... 96 Analytical procedures...... 98 Sample preparation...... 98 Metal analysis...... 98 Metal Quality Control ...... 98 Organic method: OCP extraction, clean up, analysis and confirmation of OCP ...... 98 OCP Quality Control ...... 98 Depuration kinetic rate determination ...... 99 Data treatment and statistical analyses ...... 100 Results ...... 101 Size, Total Weight and Roe Weight ...... 101 Laboratory and field depuration for cadmium, lead, nickel, vanadium ...... 103 Laboratory and field depuration for OCPs 4,4‘DDE and Dieldrin in roe ... 109 Depuration rate and Biological Half Life...... 112 Discussion ...... 114 Metals ...... 115 The organic phase ...... 118 Conclusion ...... 120 5.Bioaccumulation of metals and PAHs in Australian sea urchin Heliocidaris erythrogramma ...... 122 Abstract ...... 122 Introduction ...... 123 Methods ...... 125 Urchin Collection ...... 125 Exposure ...... 125 Spiking preparations...... 126 Experiment ...... 128 Analytical Techniques ...... 129 Metals ...... 129 PAH and Lipids ...... 129 Statistical Analysis ...... 131 Results ...... 132 Size , weights, lipid percentage ...... 132
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Metals in urchins upon collection ...... 135 Metals in roe and eggs...... 136 PAHs in roe ...... 138 Discussion ...... 140 Conclusion ...... 145 6. General Discussion & Conclusion: Summary and Recommendations ...... 146 Summary ...... 146 Comparison of two species ...... 152 Conclusion ...... 154 References ...... 157 Appendices ...... 172 Appendix 1- Abbreviations ...... 172 Appendix 2 Body parts and taxonomy of sea urchins...... 175 Body parts of the sea urchin ...... 175 Taxonomy ...... 175 Paracentrotus lividus ...... 175 Heliocidaris erythrogramma...... 175
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Chapter1
1.General Introduction
One of the biggest challenges of this century is managing for sustainable development under scenarios of exponential population growth (UNFPA et al. 2002). World population is projected to grow to 7 billion by the end of this decade (Census Bureau 2010) and it can be expected that industrial and domestic waste production will increase due to this expansion. If not managed appropriately, increased waste production will lead to more contamination of marine and terrestrial environment. Whereas, the chemical revolution of the twentieth century contributed greatly to human well-being, it also resulted in the release of toxic chemicals, some of which have persisted in the environment, travelled thousands of kilometres from where their sources, and threatened the long-term health of ecosystems (UNEP 2005). Some toxicants found their way into the food chain and were biomagnified from one trophic level to the next, causing health hazards to consumers (Gray 2002, Mizukawa et al. 2009). In fact, many toxicants have been detected lately in crops, meat, drinking water and seafood. This has led the World Health Organisation (WHO) to declare the safety of food a priority issue for humanity and this decade‘s most challenging matter (WHO 2002). Today, international food safety regulations (FDA, CODEX, European Union) are being updated towards more stringent values, which is an indication of how serious issues of food safety are taken (WHA 2010, WHO 2010).
Unfortunately, the marine environment has long been treated as a suitable location for waste disposal. Sea food can become easily contaminated because marine biota accumulates toxicants, disposed in their environment, via water or diet (Warnau et al. 1995b, Kruzynski 2004). Some marine biota, in particular, are not only an important source of food because of their high nutritional value (Burtin 2003, Willems et al. 2006) but also a source for agricultural (Garau et al. 2012) and pharmaceutical products. In fact, marine bivalves are fished for their shells, which have high calcium content and which are used for treatment of bone diseases like osteoporosis (Appleton 2000), and marine algae are extensively used in cosmetics and food industry (Burtin 2003, Ngo et al. 2011).
Marine biota, which are the subject of fisheries, often need to be tested before being put on the market. On many occasions, marine products have been rejected because toxicants in them exceed
1
1.Introduction
safety limits (Kruzynski 2004). This rejection has caused economic and reputational losses especially because of the new trend in food traceability (FAO/WHO 2002). One way of evading this situation would be to establish biomonitoring programs which may detect toxicants in biota while they are still in their natural habitat. Such programs can give authorities time to act on contamination before the damage becomes irreversible. Biomonitoring will also help minimize the rejection incident of marine products that have already been harvested. It will help avoid the potential waste of food/energy and would therefore assist in environmental protection.
Biomonitoring programs are only one part of ecotoxicological studies which can also include studies of the spatial and temporal variation in abundance of biota, exposure in the laboratory (as a mean to interpret any events in the field) and kinetics of uptake and depuration (Boudou et al. 1997). A biomonitoring program requires the identification of biota which is known to accumulate contaminants. However, for a biota to be nominated as good indicator, it must fulfil certain criteria such as being ubiquitous, abundant, easy to sample, with affinity and direct relationship to external toxic concentrations (Gerhardt 2000x, Ravera 2001).
Bioaccumulation, in aquatic organisms, is the process that causes chemicals to concentrate in the organism compared to the chemical concentration in ambient environment, water and/or sediment (Ruus et al. 2005). But for a chemical to be accumulated, it must be in a bioavailable chemical form (Boudou et al. 1997). When this bioaccumulation results in concentration levels above the food safety levels or if it reaches near toxicity, then depuration becomes necessary. Depuration is the process by which organisms are held in tanks of clean seawater under conditions which maximize the natural elimination activity which results in the expulsion of toxicants from tissues (Lee et al. 2008). Depuration studies in particular have important repercussions especially if levels of the toxicants, which are not regulated by the organism, reach a toxic level and must be purged with assistance of chelating agents.
While many developed countries such as some states of the European Union, United States of America and Australia have already set up biomonitoring programs in the 1970s, other developing countries (e.g. Lebanon) lack these programs for either lack of internal legislation or lack of funds. With regard to internal legislation, this has been corrected by the United Nations when it set up few international protocols to monitor and protect the environment (e.g. the Stockholm protocol). More than 150 countries signed the protocol which was initiated on 17th of May 2004 (UNEP 2001, 2005). However, for developing countries, research funds are probably still an issue.
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1.Introduction
Lebanon is a country where environmental research was still rare up to 2006. This country lies on the eastern side of the Mediterranean Sea which itself is subject to many wastes being discharged daily without any treatment (UNEP 2002). Lebanon has been ravaged by war for the last 35 years and has not upgraded its infrastructure since then , i.e waste water treatment plants have not been set up yet (World et al. 2003). Lebanon also lacks serious biomonitoring programs (IMO/REMPEC/UNEP-MAP 2006, UNEP 2007) despite being a signatory to the Stockholm convention, which requires all its participants to monitor for Persistent Organic Pollutants (POPs). Meanwhile, seafood is heavily consumed in Lebanon without prior testing. In addition, fishing related rules and regulations have not been updated in the last 40 years. Interest in research and biomonitoring in general, has risen recently in Lebanon following the 2006 events when an oil spill of more than 15,000 metric tonnes was discharged along ca 70% of its coast (UNEP 2007, Coppini et al. 2010).
Oil spills can occur in any place in the world, be it on land or in water and are not restricted to developing countries (IMO/REMPEC/UNEP-MAP 2006, Amiard et al. 2004). Australia‘s latest oil spill occurred in 2010 where four metric tonnes of oil were discharged in Keppel island Queensland (AMSA 2011). Many organisms native to Australia are good bioindicators and have been used in biomonitoring programs (Smith et al. 2003, Roberts et al. 2008). As some fisheries are on the brink of collapse and recent interest in some new fisheries emerges in Australia (Fox et al. 2000), this interest will prompt authorities to biomonitor for toxicants in new species. Particular interest should be placed in monitoring of toxicants stemming from oil spills (e.g. nickel, vanadium, polycyclic aromatic hydrocarbons (PAHs) to mention few) especially among benthic animals, which live in close proximity to the sea floor.
Nickel and vanadium are metals that can be found in crude or refined oil (Tronczynski et al. 2004). In the Erika oil spill, their concentrations in marine organisms, was used to trace the organisms‘ exposure to oil deposits along the coast. (Amiard et al. 2004). Cadmium and lead can also be found in refined oil but their presence is mostly monitored due to their severe health effects on biota (ATSDR 2007, 2008). PAHs are main components of crude oil for which a range of biological effects have been demonstrated, including: acute toxicity, carcinogenicity, mutagenicity, teratogenicity (ATSDR 1995).
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1.Introduction
Echinoderms are benthic animals which have been extensively used in ecotoxicological studies (Coteur et al. 2003, Sugni et al. 2007). Some echinoderm species, particularly urchins and sea cucumbers, are consumed by humans worldwide. Echinoderms have certain developmental features that are similar to humans. In fact, 70% of the sea urchin‘s genes resemble those of a human‘s, and it is believed that there are similarities between the immune response systems of sea urchins and humans as well (Cervello et al. 1996, Coteur et al. 2001, Sea Urchin Genome Sequencing Consortium,et al. 2006). Studies have shown that sea urchins do accumulate metals and bacteria (Portocali et al. 1996, Warnau et al. 1998). In the Mediterranean Sea, Paracentrotus lividus is a dominant urchin (Guidetti et al. 2003b, Guidetti 2004, Guidetti et al. 2005). In species such as Paracentrotus lividus, few studies looked at the bioaccumulation of different organic subclasses especially in the occurrence of an oil spill (Danis et al. 2005). P. lividus is a suitable organism to be used as a bioindicator as observed by many researchers, and I hypothesize that it could accumulate all classes of toxicants present all along the Lebanese coast. The fact that this organism is a grazer rather than a filter feeder will allow me to get a better and wider picture of the contamination in the field (Bulleri et al. 1999). So sea urchins, specifically P. lividus , have been used as indicators of bacteriological and metallic contamination as previously stated,. But to the best of my knowledge, no prior studies were done on the organic contaminants e.g. OCPs and PAHs, some of which are suspected to cause endocrine disruption (ATSDR 1995, 2002b). Moreover, no studies have been done on a mixture of toxicants to account for any additive, synergistic or antagonistic effects on bioacumulation. Extensive research on P. lividus has also shown that its larvae is sensitive to many ranges of chemical toxicants (Radenac et al. 2001, Pesando et al. 2003, Pesando et al. 2004).
In Australia, many sea urchins exist but one urchin in particular, Heliocidaris erythrogramma, a native to Australia, is being considered for a new fishery in New South Wales (Fox et al. 2000). Before considering this organism as a new potential seafood, a pilot study aiming at setting up toxicant baseline values of H. erythrogramma is recommended. Furthermore, testing the ability of this urchin to accumulate toxicants, which marks oil spills, is recommended.
Aims of study The aims of this study were:
a) To begin a biomonitoring program in Lebanon for purposes of protecting the environment and protecting human consumers and seafood.
b) To assess the suitability of Paracentrotus lividus as a bioaccumulator species through laboratory bioaccumulation studies.
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1.Introduction
c) To compare, in sea urchins, the depuration of common oil spill contaminants in the laboratory vs. the field.
d) To check baseline levels in Heliocidaris erythrogramma of metals associated with oil spills and to test the ability of H.erythrogramma to bioaccumulate oil spill toxicants.
This research, therefore, will hopefully be an important contribution to scientific knowledge in general and to the country of Lebanon specifically. The uptake of toxicants, measurements of bioconcentration factor (BCF), toxicity and depuration of P. lividus will provide some data that can be vital to the conservation and management of marine ecosystems and to the aquaculture industry of sea urchins. It will add to our understanding of P. lividus’ physiology when this sea urchin is exposed to a variety of toxicants not all previously tested in this organism before to the best of my knowledge.
This research fits within the framework of Med Pol-MAP (the marine pollution assessment and control component of Mediterranean Action Plan). MED POL assists Mediterranean countries in the formulation and implementation of pollution monitoring programmes, including pollution control measures and the drafting of action plans aiming to eliminate pollution from land-based sources as required by the Barcelona convention (UNEP 1976).
During the course of this study the country of Lebanon was subjected to armed attack and a major power plant (Jieh Power Plant) was bombed in mid July 2006. This resulted in a large oil spill (ca 15,000 tonne) that spread north along the coast of Lebanon. During this incident and the armed conflict there was a complete sea blockade along the coast of Lebanon that lasted for 7 months (from July 2006 - January 2007). These events affected the course of the study in terms of damage to the research laboratory and electricity supply problems for complex analytical instruments but more importantly in terms of the potential effect on the biology of the study organism. Following the oil spill, there was an increased interest in marine contaminant issues within coastal Lebanon. The United Nations (UN) sent a post-conflict assessment team to investigate the impact of the conflict on the environment. Oysters were collected, among many other matrices, for polyaromatic hydrocarbons analyses. The UN reported that Lebanon indeed lacks appropriate infrastructure for contamination management and monitoring programs (IMO/REMPEC/UNEP-MAP 2006). So if the results of this study contribute to the establishment of a more permanent biomonitoring program, Lebanon would then be both abiding by international protocols (Barcelona and Stockholm) and watching over the safety of both the marine ecosystem and the consumer health.
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1.Introduction
Thesis outline This thesis is structured as a series of stand-alone manuscripts as most chapters are in the process of being prepared for submissions to peer-reviewed journals (after some minor modifications like adding names of authors etc.). Therefore, there may be some slight repetition between chapters. Below is an outline of the content of each chapter.
Chapter 2 describes a field survey (temporal variation over five years) and assesses the biomass, density and size of sea urchins P. lividus relative to its distance from a potentially polluting source. It looks at the bioaccumulation in the body parts of the sea urchins relative to the distance from the polluting source. A body burden was determined for the year 2007 where the maximum density of urchins was found. Sea urchins from another site to the north of the coast were also checked in 2007 for density, biomass, size, roe weight as well as metal accumulation. A literature review, on abundance, biomass and metal concentrations of sea urchins for the last thirty years in the Mediterranean basin was done, and the data from this study was compared to the literature review data from different geographical spots in the Mediterranean Sea.
Chapter 3 describes the laboratory-exposure of sea urchins to a cocktail of toxicants (organic, inorganic and bacterial) to mimic real environmental conditions where these organisms are subject to a mixture of infinite number of toxicants. The research attempts to shed some light on differences in bioaccumulation between different body parts as seen in the field. Bioconcentration factors were determined for all toxicants used in the exposure experiment.
Chapter 4 assesses the depuration rate in the roe of the sea urchin P. lividus . Depuration of four metals and two organochlorinated pesticides (OCPs) was undertaken in two locations (1) in the field by translocation and (2) in the laboratory assisted with the addition of a chelating agent. This chapter concludes the studies on P. lividus.
Chapter 5 examines the bioaccumulation of four metals and three PAHs in Heliocidaris erythrogramma, a species suggested for fisheries in Australia, and compares differences in bioaccumulation between males and females. It also compares the baseline metal concentrations to the safety limits in food.
Chapter 6 provides a synthesis of the main findings, a general discussion with comparison between the two species of sea urchins, a discussion of the implications of this research and recommendations (in italic) for each chapter. Topics for future research arising from this work are also discussed.
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Chapter 2
2.Temporal variation in abundance, biomass, size and metal content of the sea urchin Paracentrotus lividus in the South of Lebanon
Abstract
Benthic animals are potentially good indicators of contamination. The sea urchin, Paracentrotus lividus, was used as a potential biomonitor of a potentially contaminated site off the Lebanese coast. Strategic surveys were undertaken for five years to assess whether proximity to a major source of pollution affected abundance, biomass, size and metal content of the urchin. Metal concentrations in four body parts of the sea urchin did not vary with distance from the major pollution source. Size, abundance and biomass also did not differ with distance from the potential pollution source and were more directly affected by an unplanned closure to the fishery in 2007. Only during this year, another location (Tripoli), situated on the northern coast, was censused for sea urchins‘ abundance, size, gonad index and metal content. This location is a marine protected area but is subject to contamination. The marine protected site had higher abundance, biomass, gonad index and metal loads than the Sidon site in 2007. In Sidon, the size of sea urchins did not differ between sites or years. The metal concentrations of roe exceeded the EU food safety guidelines. Urchin density and biomass were low but within the range observed for the Mediterranean. Sea urchins, collected from Sidon in 2008, had significantly higher metals concentration than sea urchins collected in 2007. The distribution and metal content of the urchin is not considered a sensitive indication of contamination, however, the contaminant loads measured are considered as a direct concern with regards to human health, as consumption of urchin roe is currently unregulated.
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Chapter 2.Temporal variation
Introduction
Marine ecosystem contamination is a serious problem in many coastal areas of the world (Tihansky 1973, UNEP 1990, Anderson et al. 2002). Contaminated effluent and runoffs enter coastal waters and may cause mortality of commercially and ecologically important organisms or result in the bioaccumulation of toxic substances (Bennett et al. 2001, Pulido et al. 2003). This may in turn impact human communities that are reliant on seafood production and consumption. In fact, sea food products are often rejected if values of toxicants are found to exceed regulations (Kruzynski 2004). Additionally contamination may result in further damage if affected organisms are important functional components of the ecosystem. Whilst substantial literature exists on the contamination of seawater (Manfra et al. 2005) and seafood (NSW-HD 2006) in some highly industrialised countries, other developing countries are not so well studied. The paucity of baseline data in developing nations may result in policy and risk assessment procedures that may be deemed inadequate with regards to the protection of both human and ecosystem health. Metals are a ubiquitous contamination problem in coastal areas. They are highly persistent in the environment and can be toxic at very low concentrations (Kljakovic´-Gašpic´ et al. 2010). Accumulation of metals in marine organisms can occur at levels much greater than ambient environmental concentrations (Gray 2002, Cebrian et al. 2003, Karouna-Renier et al. 2007). Several metals are also poisonous to humans who consume metal contaminated food (Albanese et al. 2008 ). Metals, such as cadmium and lead can cause physiological perturbations, growth suppression and reproductive damage in marine organisms and in humans (ATSDR 2007, 2008, Albanese et al. 2008 , García-Pérez et al. 2010).
Many marine organisms used in marine bioaccumulation studies, such as mussels, oysters, crabs, sea urchins and fish, are dominant members of coastal communities (Boisson et al. 1998). These organisms are also targeted for human consumption (Karouna-Renier et al. 2007, Julshamn et al. 2008). Sea urchins in particular are important ecosystem engineers (Hereu et al. 2004, Tomas et al. 2004) and have been shown to accumulate a variety of toxicants in all body parts (roe, test, spine and teeth, appendix 2) (Augier et al. 1995, Portocali et al. 1996, Pegano et al. 2002, Danis et al. 2005).
In many parts of the world the sea urchin is considered a delicacy and the roe is consumed in large quantities because of its high nutritional value (Liyana-Pathiranaa et al. 2002, Boudouresque et al. 2007). This is largely the case in the Mediterranean Sea where the abundance of sea urchins is simultaneously impacted by both pollution and over-harvesting (Sala et al. 1998, Steneck 1998).
The Mediterranean is heavily influenced by contamination which may directly impact marine invertebrate communities (UNEP 2002). In addition to this problem, many fisheries are on the brink of collapse due to over-harvesting (Garcia-Charton et al. 2008). This results in a situation in
8
Chapter 2.Temporal variation
which both overharvesting (Gianguzza et al. 2006) and contamination may be reducing catch sizes; while the consumption of contaminated food may be impacting human health. Many studies have looked specifically at the ecology and fishery of the rock sea urchin Paracentrotus lividus (Waker et al. 2001, Guidetti et al. 2003a, Guidetti et al. 2003b, Guidetti 2004, Boudouresque et al. 2007). This sea urchin is a common benthic grazer on rocky substrata and is distributed throughout the Mediterranean Sea. It is also present in the North-East Atlantic, from Scotland and Ireland to Southern Morocco and the Canary Islands, including the Azores (Boudouresque et al. 2007). The P. lividus is an echinoderm which feeds on algae, seaweeds, seagrasses and is restricted to shallow waters where it shelters in rock or crevices (Sala et al. 1996). The disappearance of this particular sea urchin has consequences for other marine organisms (Tomas et al. 2004). Changes in their relative abundance can dramatically change community composition and structure (Lozano et al. 1995, McClanahan et al. 1997, Palacin et al. 1998, Tomas et al. 2004). Throughout the Mediterranean basin, the sea urchin has been the subject of intensive studies for metal bioaccumulation. They have been shown to accumulate metals in all body compartments, with levels above ambient water concentrations (Auernheimer et al. 1997, Warnau et al. 1998, Soualili et al. 2008). Unfortunately, monitoring data are lacking in many regions of the Mediterranean, in particular the eastern coast.
Lebanon‘s coastline, which is situated on the eastern side of the Mediterranean Sea, is a 225 km stretch of shore. It is subject to constant environmental disturbances as a result of wars, non- functional state regulatory bodies and dysfunctional infrastructure including solid waste recycling and sewage treatment plants (World et al. 2003, UNEP 2007). Being a signatory of several international conventions for the protection of the Mediterranean Sea from environmental pollution, Lebanon has an obligation to initiate monitoring programs, which are essential tools for ecological risk assessment. Currently, assessments of spatial or temporal variation of benthic communities in Southern Lebanon are scarce. Anecdotally, it has been conjectured that pollution and over fishing in Sidon's harbour are causing a drastic decrease in fauna and flora of Lebanon especially sea urchins (Pers.comm.). Nevertheless, ecological data is necessary to support or refute such claims.
The present study used the sea urchin Paracentrotus lividus as a model organism to assess effects of coastal pollution on benthic organisms. Bioaccumulation of four metals in various body parts of the urchin and possible relationship between contaminant loads and distance from a major coastal pollution source were evaluated. The present study also investigated temporal variation in the density and biomass of sea urchins with distance from the pollution source from 2005 through 2009.
9
Chapter 2.Temporal variation
I tested the following hypotheses:
1. Sea urchin tissue contamination decreases with increasing distance from the major coastal contamination source 2. Sea urchins carry the same metal loading in each of their four major body compartments (roe, test, spines teeth). 3. The density of sea urchins decreases with increasing proximity to the major coastal contamination source. 4. Sea urchin density is temporally stable.
Material and Methods:
Survey Methods and Design
For the purposes of the temporal variability in abundance study, sea urchins (Paracentrotus lividus) were censused and collected at multiple sites along a 900 m stretch of the Lebanese coast, 200 m offshore and in close proximity to Sidon Harbour, South of Lebanon (Figure 1). This location is one of the most severely contaminated areas of the Lebanese coast (GreenPeace et al. 1997). Over a thirty year period, industrial, domestic and agricultural waste have been deposited on the beach by the municipality (Abouzeid 2003). The waste has accumulated to form a mountainous structure spanning approximately 200 m in length, 20 m in height, and 40 m in width. The effluent, released from the (waste deposit) mountain, spreads along the sea floor and is carried northwards by the prevailing southerly surface current (Boulos 1996). The effluent and litter are posited to create an almost sterile area opposite the discharge point (no biota seen when diving, personal observation). The conditions are generally worse during winter as ―storms‖ push tons of garbage into the sea (Zaatari 2007b). To date, there has been no effort to determine the impact of this toxicant cocktail that is discharged daily, without treatment, into the marine ecosystem.
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Chapter 2.Temporal variation
c 750 m distance
Mediterranean sea 600 m distance
Sidon‘s islands 450 m distance
10m 50m transect line with 5 plots per line with 300 m distance 10m distance between 2 plots and 5 m distance
between two transects lines Current Sidon Harbour direction
Mountain of waste deposit
Figure 1 Sampling location: a- the Mediterranean sea,. b- map of Lebanon featuring Sidon‘s Harbour, South of Lebanon where waste of municipal and industrial solid deposit is considered a major source of pollution and Tripoli, a marine protected area, North of Lebanon. c- Sampling plan off Sidon‘s harbour with 4 sets of 3 transect lines at 300 m, 450 m, 600 m and 750 m from the source of pollution.
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Chapter 2.Temporal variation
Surveys were conducted in the month of May (start of spawning season) in each year from 2005- 2009 (five census events). The survey design featured multiple transects at four distances from the major source of coastal pollution; the waste deposit mountain (Coordinates:N 33 32 971 E 03521.793 and N 33 33232 E 035 21.766). The first three transect lines were placed 300 m from the garbage mountain. This was the minimum distance divers were prepared to enter the water for the purposes of environmental research and for health reasons. The three transects were repeated at 300, 450, 600 and 750 m away from source of pollution. At each distance, a 10 m weighted rope was thrown from the boat perpendicular to the coast. Three 50 m long transect lines were each attached to the 10 m rope at its beginning, its middle and its end (each separated by 5 m) and thrown parallel to the beach in the direction of the current (Figure 1c). On each 50 m transect line, a mark was put every 10 m. A square meter grid (1 m x1 m) was thrown randomly between these marks. This is the plot quadrant. One diver was assigned to each transect. The diver counted, sized and collected sea urchins from each plot. Canvas bags containing sterilised acid washed containers were attached to the middle transect for water collection. Divers were given yellow plastic sheets superimposed with a clear plastic urchin size chart. The urchin size chart was provided for each diver to ensure consistency in sizing. The chart displayed three different sizes of urchins (Small:< 2cm, Medium: 2-5cm, Large:> 5cm). The depth at which sea urchins were collected, along transects, ranged from 5 m to a maximum of 12 m. The diver took a maximum of three sea urchins of uniform size from each quadrant in the years 2007 and 2008 and placed them in a labelled zip lock bag. The zip lock bag was placed in a canvas bag attached to the nearest line mark. A minimum of three sea urchins from each quadrant were necessary to conduct toxicant analyses.
I collected more sea urchins off Sidon‘s coast in 2007 than in the two previous years (10 fold more). There was now a sufficient number of urchins for the tissue analysis. It was still not clear if I would be able to collect sufficient urchins for metals analysis in any other year from Sidon and hence I decided to sample urchins for a comparison from a back-up reference location in 2007. I looked for another location where fishing was forbidden but contamination of seawater was substantial. Tripoli, a city lying North of Lebanon, is host to a Marine Protected Area extending from the coast to the three islands that was established in 1980 and encompasses 4,200,000 m2. Fishing is prohibited in the Marine Protected Area of Tripoli but there are outlets from which industrial and domestic wastes are discharged directly into the harbour to the south of the marine protected area. I applied the same protocol census and collection used in Sidon but on one site only (3 transect lines). My census point in Tripoli was 600 m from the main effluent outlet in the marine park. So for the comparison of abundance, biomass, roe weight, gonad index and metals concentrations between the two locations (Sidon and Tripoli) for 2007, I took one site
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Chapter 2.Temporal variation
from Sidon (600 m from the rubbish mountain, 3 transect lines) and compared it to one site in Tripoli (also 600 m from a polluting source, 3 transect lines).
Tissue contamination measurements Sea urchins were immediately brought to an ISO 17025 accredited, analytical laboratory (Environment Core Laboratory) at the American University of Beirut and treated as follows: sea urchins were washed on the exterior with de-ionized water before being weighed on an analytical balance. The diameter of each sea urchin was measured with a Vernier caliper then recorded to the nearest 1 mm. The colour of the sea urchin was also noted.
Spines were stripped off the test- body and both spines and body were washed again with 1% nitric acid for few seconds to remove adsorbing metals (Du et al. 2011) then rinsed with de- ionized water. Leaching of metals was not observed but if it did occur, it would not have altered the ratio of metals in the different body parts of the sea urchin. Each sea urchin was dissected and teeth were separated from the intestines and washed with 1% nitric acid. Excised sea urchin parts (spines, teeth and body-test, Appendix 2) were dried for 48 h at 65oC in a forced air oven, blended and homogenized separately. Roe was weighed first then washed with 1% nitric Acid (ICP-MS grade), de-ionized water and freeze dried. Approximately 0.1 g to 0.5 g (dry weight) of each of the body parts was digested with 3 ml of a 70% nitric acid (ICP-MS grade, purchased from Fisher) in a closed vessel microwave digester for 20 minutes at 1000 W. For roe, 2 ml of 30% hydrogen peroxide (ICP-MS grade, purchased from Fisher) were added to the 3 ml of nitric acid to ensure full digestion of organic matter. The digests were transferred to acid washed vials and diluted with deionised water to appropriate volumes. When the amount of metal in the samples exceeded the highest point of calibration curve, additional dilutions were performed for both samples and quality control standards.
The quality control protocol used in the microwave digestion consisted of the analysis of a blank, a standard reference material (NIST oyster tissue 1566b), duplicates and spikes. To check for errors during sample preparation, 10% of randomly chosen samples were re-digested. Digests were run on an ion coupled plasma-mass spectrometer (ICP-MS 7500ce-Agilent equipped with a Cell Dynamic Range and an autosampler) for cadmium (Cd), lead (Pb), nickel (Ni) and vanadium (V) determination. The first two metals were chosen based on their toxic effect on reproduction of sea urchins and because they are bi-products of industrial activities (Au et al. 2003, Xu et al. 2010). Vanadium and nickel were chosen in light of a recent oil spill in Lebanon (Coppini et al. 2010) as these two metals are usually found in dispersed fuel oil (Tronczynski et al. 2004). The quality control protocol used when analysing samples on ICP-MS consisted of the following: a tune up of instrument, a blank of system, a calibration standard curve of the four metals (a standard curve with a correlation of >0.995), a quality control check every 10 samples (standard
13
Chapter 2.Temporal variation
purchased from two different suppliers), a quality control for the matrix interference, and a mini calibration curve every twenty samples (lowest value, mid value and upper value of curve). An internal standard mixture consisting of scandium, thallium, ytterbium and cerium (purchased from Agilent) was added to all samples and all calibration levels. Recoveries for all quality control samples were accepted when values were within the control limit charts of the corresponding metals (ca 80-120%). The generated data was treated for both the digestion and analyses and calculated from both the external standard curve and the ratio of internal standard (IS) using Chemstation (B.03.07). Further data treatment was applied on the cadmium and lead in the roe (25% of data) when bias was more than 20% and results were expressed as microgram of metal per g of dry weight (µg g-1). Analytes‘ concentrations were corrected for extraction efficiency by multiplying the average recovery of the surrogate/internal standards by each analyte concentration detected in the sample. For statistical purposes values less than the detection limit were reported as half the limit of quantification, which was 1 ug l-1 for water samples and 0.01 ug g-1 for tissues. All methods used for digestion and analysis were sourced from internationally approved methods with some modifications (EPA-USA-200.8 1994, EPA-USA-3052 1996, APHA 1999). In order to balance the data set when comparing Tripoli and Saida,(in terms of comparing equivalent number of samples to be used in metal analysis) I randomly chose a subset of samples from Sidon in 2007 and an equal number of transects for biomass, size, roe weight and gonad index determination.
All glassware, plastic ware and utensils used for analytical measurements were soaked and washed with 1% nitric acid (ICP-MS grade), rinsed with deionised water then left to dry in a dust free room.
Statistical Methods Sufficient urchins for toxicant analysis were only available from all distances in 2007 (this sampling event followed a military blockade of the coast that effectively removed fishing pressure). The relationship between distance from pollution source and amount of metal accumulated in each of the sea urchin body parts was tested using a Two Factor ANOVA with the two fixed factors of distance (300, 450, 600 or 750 m) and body part (roe, test, spines or teeth). Sufficient urchins for analysis were collected in 2008 from the 600 m distance category only. A Two Factor ANOVA (α=0.05) was used to test for differences of bioaccumulation in body parts of sea urchin for two consecutive years ( 2007 and 2008) at the 600 m distance only.
I was not able to contrast for differences in density, biomass and size of sea urchins relative to the distance from source of pollution across years using Two Factor ANOVA as many plots in some sites and in some years contained zero values. Instead, the mean and 95% confidence intervals were plotted, and an overlap in confidence intervals between sites or years was considered as not
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Chapter 2.Temporal variation
significantly different. Values were first averaged in each transect and presented for each site in each respective year. Values were also presented as the average of all sites for each year. I also analysed the data using permutational analysis of variance (PERMANOVA) that does not assume homogeneity of variances. Our results from the PERMANOVA (using PRIMER v6) (Clarke et al. 2001, Clarke et al. 2006) were consistent with the interpretation of confidence intervals and hence I present only the graphical confidence intervals for the sake of brevity.
To test for differences in gonad index (wet weight of roe X 100/total weight of organism), roe weight, biomass and size between sea urchins collected from Tripoli and Sidon, a student ―t‖ test was used with location as factor (Sidon South, Tripoli North) and to test for difference of density for one distance (600 m distance from pollution source) between two locations (Sidon, Tripoli). A ―t‖ test was also used to detect significant differences in levels of sea water metals (Ni,V, Cd and Pb) across the years averaging values from the four distances (300,450, 600 and 750 m).
Assumptions of normality and homogeneity of variances were examined using the Shapiro-Wilk and Levene tests respectively ( Quinn et al. 2002). When the assumptions of ANOVA were not met, non-normal data sets were inverse, log or square root transformed. Tukey‘s Studentized Range test was the multiple comparison procedure used in conjunction with the ANOVA to determine which means were significantly different from others. Results of post hoc tests are appended to the ANOVA table. If after transformation, data set were still not normal, differences between group means were assessed using the non parametric Kruskal-Wallis One Way Analysis of Variance (for each factor) on Ranks with a non parametric Student Newman-Keuls (SNK) for post-hoc comparisons (in this case no interaction between factors was considered).Statistical analyses were performed with the SPSS (2010) version 18.
Comparison with published reports A literature review was conducted by combining the search terms ‗P. lividus‘ with the term ‗metal‘, and ‗P. lividus‘ with the terms ―density‖, ―size‖ and ‗biomass‘ in scientific databases. References reporting metal concentrations in body parts of P.lividus from field sites around the Mediterranean as well as studies reporting density, size and/or biomass were selected for review. Reference lists of articles identified in this preliminary search were also examined to capture studies published prior to 1990. Since this is the first baseline study to be done on sea urchins in Lebanon, the findings were used to assess the status of the sea urchin P. lividus relative to the status of P. lividus in other locations of the Mediterrranean.
I also compared the levels of some chemical and physical parameters in Sidon‘s seawater to international guidelines especially recreational values and trigger values of ANZECC (2000a, b). According to ANZECC (vol. 1), ―the guidelines provided are designed to help users assess
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Chapter 2.Temporal variation
whether the water quality of a water resource is good enough to allow it to be used for humans, food production or aquatic ecosystems‖. In case water quality does not meet the water quality guidelines, management action could be triggered to remedy the problem. For metals, the comparison was made between levels of metals in Sidon‘s area for the last 5 years and the background levels of metals in the Mediterranean‘s coastal and open water, Australia and the rest of the world.
Results
Density, biomass and size
Sidon 2005-2009 Densities of sea urchins varied dramatically amongst years but there were never any urchins recorded within 300 m of the major pollution source (Figures 2-a1 and a2). Average urchin density ranged from 0.03 m-2 in 2009 to 1.55 individual m-2 in 2007. Ninety-five percentile intervals indicated that urchin densities were greater than zero at 600 m distance in years 2006, 2007 and 2008. However, presence/absence data indicates that restricted number of urchins were regularly present at the 450 transect locations. I also recorded very low densities of urchins at the 750 m transect plots in each year except 2007, when densities were much higher.
The average Biomass (sum of total wet weight m-1) did not differ across distances from the pollution source except for the 600m distance in 2007. Across years, it ranged from 1.8 to 30.0 g wet weight m-1 with the years 2006 and 2009 scoring the lowest values (Figure 2-b1 and b2 ).
Examination of 95 % confidence intervals indicate that size (diameter without spines in cm) of urchins were biggest in 2008 but did not differ with distance from the pollution source (Figures 2- c1 and c2). The size of the sea urchin (P. lividus) (without spines) ranged from 3.3 cm to 3.8 cm across distances from polluting source and for the period of the study.
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Chapter 2.Temporal variation
Figure 2 a-1 Density (number of individual m-2 ±95% Confidence Interval (CI),b-1 biomass (wet weight in g m-2±CI ) and c-1 size (diameter without spines in cm±CI) of sea urchin Paracentrotus lividus in Sidon, South of Lebanon averaging sites for each year; a-2,b-2 and c-2 represent density, biomass and size respectively for transects averaged for each of the four distances and plotted across the years.
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Chapter 2.Temporal variation
Sidon-Tripoli Comparison 2007 When comparing the density, biomass and size between two locations (Tripoli & Sidon) in year 2007, density and biomass in Tripoli were triple those of Sidon (Figure 3 a,b). Roe wet weight (Figure 3d) and the gonad Index (Wet weight of roe x100/ total wet weight of sea urchin) (Figure 3e) were higher in the sea urchins of Tripoli than in the sea urchins of Sidon. However, there was no difference in size between the two locations (Figure 3 c).
Figure 3 a-Density (number of individual m-2 ±SE),b- biomass (wet weight in g m-2±SE ), c- size (diameter without spines in cm ±SE), d- roe weight (wet weight in g m-2±SE) and e- Gonad Index (GI, wet weight of roe in g x 100/ total wet weight ±SE) of sea urchin Paracentrotus lividus in Sidon, South of Lebanon and in Tripoli North of Lebanon for the year 2007. Results of ―t‖ test are enclosed (significance <0.05).
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Chapter 2.Temporal variation
Metal bioaccumulation in body parts
Sidon 2007 Sufficient urchins were available for tissue analysis from three distances (450, 600 and 750m) in the year 2007. In samples from this year, measurable concentrations of all elements studied were found for all body parts in the sea urchin P. lividus. The mean metal concentrations (µg g-1 ±SE) of V, Ni, and Cd did not vary with regards to distance from the pollution source for any body part of the urchin (Figure 4 &Table 1). The exception was the concentration of Pb, which increased with distance from the source, although this appears to be largely driven by concentrations found in the roe only (Figure 4d).
Tissue concentrations of vanadium differed among body parts (Figure 4a). In general there was greater concentration of V in the roe. The lowest concentrations were found in the spines. The spines were approximately 10 times lower than those found in the roe.
The analysis of variance for nickel revealed a wide fluctuation in accumulation among body parts (Figure 4 b & Table1). No interaction was detected between body parts and distances. The highest mean for nickel concentration was in the spines (1.564 to 2.058 µg g-1) and the lowest in the roe (0.209 to 0.243 µg g-1 dry wt) as opposed to the vanadium measurement. This was consistent throughout all the sites.
The ANOVA for cadmium showed a different pattern again from the vanadium and nickel. Body parts differed in their concentrations at the 450 and 750m distances (Figure 4 c & Table1). Similarly to vanadium and nickel, no interaction was detected between body parts and distances. The mean cadmium concentration in the test was the highest and ranged from 0.146 to 0.218 µg g-1 dry weight whereas concentrations of cadmium in spines, roe (except at 450m distance) and teeth were not different from each other.
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Chapter 2.Temporal variation
Figure 4 Metal bioaccumulation in µg g -1(±SE) of sea urchin P. lividus’ body parts in Sidon, South of Lebanon for the year 2007 across 450, 600 and 750 m distances from major pollution source (a-Vanadium, b-Nickel ,c-Cadmium and d-Lead).
Lead concentrations differed according to body parts and distance although there was no interaction (Fig 4 d & Table1). The mean lead concentration in the roe increased with increasing distance from pollution site. It ranged from 0.082 to 0.37 µg g-1 dry wt. The lowest concentration was in the spines (0.002 µg g-1 dry wt which is below quantification limit 0.01 µg -1 but above detection limit to 0.001 µg -1). In the post hoc test performed on means of lead with relation to distances, concentrations were significantly lower at 450m than at 600 and 750 m.
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Chapter 2.Temporal variation
Table 1 Two Factor ANOVA testing for differences in bioaccumulation of four metals in different body parts of P.lividus related to distance from major polluting source off Sidon's harbour, South of Lebanon, in 2007. The multiple comparison tests of Tukey, performed after Two Factor ANOVA, are summarized and appended to the table. Mean metal concentrations are ranked from the left to the right by decreasing order. Terms in underlined compartments, periods or locations are not significantly different (αTukey= 0.05).
Vanadium Nickel Cadmium Lead
Data transformation Log Log Log Square root
Levene test 0.054 0.062 0.054 0.091
df MS P df MS p df MS p df MS p
Body parts 3 4.4 0.014 3 0.7554 0.000 3 0.12082 0.0144 3 44.014 0.000
Distance 2 0. 1 0.8907 2 0.0131 0.3383 2 0.0032 0.8907 2 5.543 0.000
Body parts x Distance 6 1.8 0.1464 6 0.02221 0.1386 6 0.04935 0.1464 6 0.155 0.657
Error 16 0.027 -- 22 0.011 -- 22 0.027 -- 20 0.235 --
Parameters Body parts Distances in m
Vanadium Roe Teeth Test Spines 750 600 450
Nickel Spines Test Teeth Roe 450 750 600
Cadmium Test Spines Roe Teeth 450 600 750
Lead Roe Test Teeth Spines 600 750 450
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Chapter 2.Temporal variation
Sidon 2007 -2008 Tissue concentrations of all metals differed between years (Figure 5 & Table 2). In general, there was was greater tissue concentration in 2008 than in 2007 (at least threefold).
Table 2 Two Factor ANOVA testing for differences in bioaccumulation of four metals in different body parts of sea urchin Paracentrotus lividus in years 2007 and 2008, off Sidon's harbour, South of Lebanon. The multiple comparison tests of Tukey performed after 2 factor ANOVA are summarized below the table. Mean metal concentrations are ranked from the left to the right by decreasing order. Terms in underlined compartments, body parts or locations are not significantly different (αTukey= 0.05). Vanadium was done using a non-parametric Kruskal Wallis test with each factor (year and body part) being analysed separately.
Vanadium Nickel Cadmium Lead
Data transformation Not applicable Log Log Log
Levene test Not applicable 0.087 0.272 0.078
df MS p df MS p df MS p d MS p
Body parts -- --- 0.006 3 0.37 0.000 3 0.023 0.59 3 56.91 0.000
Years -- --- 0.008 1 6.35 0.000 1 8.358 0.000 1 356.56 0.000
Body parts x Years ------3 0.45 0.000 3 0.101 0.090 3 179.75 0.000
Error ------16 0.03 -- 14 0.035 -- 15 0.009 --
Parameters Body parts Years
Vanadium Test Roe Teeth Spines 2008 2007
Nickel Test Spines teeth Roe 2008 2007
Cadmium Test Teeth Spines Roe 2008 2007
Lead Test Teeth Roe Spines 2008 2007
Using a non-parametric test for vanadium, I found difference between years and among body parts (p=0.006). The highest concentration of vanadium was in the test, followed by the roe,
22
Chapter 2.Temporal variation the teeth and the spines scoring the lowest concentration (Figure 5 a). Nickel accumulated in the test then in decreasing order in the spines and teeth with the roe accumulating the least (Figure 5 b). There was an interaction between body part and year for nickel and lead. In 2007, nickel accumulated more in spines than in the test but in the following year, this increase in bioaccumulation shifted from spines to test. Cadmium and lead accumulated mostly in the test than in teeth and spines with the least bioaccumulation occurring in the roe (Figure 5 c and d).
Figure 5 Metal bioaccumulation in µg of metal g-1 (±SE) of sea urchin P. lividus’ body parts in years 2007 and 2008 at a distance of 600 m from a major pollution source in Sidon, South of Lebanon (a- Vanadium, b-Nickel ,c-Cadmium and d-Lead).
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Chapter 2.Temporal variation
Sidon and Tripoli 2007 Sea urchins from the Tripoli site showed a much higher bioaccumulation in their body parts than sea urchins from the Sidon site (Table 3). Almost all concentrations of metals in body parts were higher in urchins from Tripoli than urchins from Sidon (Figure 6). There was also a statistical interaction between locations and body parts for all metals.
Table 3 Two Factor ANOVA testing for differences in bioaccumulation of four metals in different body parts of 12 Paracentrotus lividus. 12 sea urchins were collected from each of the two locations Sidon in the South of Lebanon and Tripoli in the North of Lebanon in the year 2007 both at a distance 0f 600 m from a major pollution source. The multiple comparison tests of Tukey, performed after Two Factor ANOVA, are summarized below the table. Mean metal concentrations are ranked from left to right by decreasing order. Terms in underlined compartments, body parts or locations are not significantly different (αTukey= 0.05).
Nickel Vanadium Cadmium Lead
Data transformation Square root Log Log Log
Leven test 0.175 0.055 0.23 0.187
df MS p df MS p df MS p df MS p
Locations 1 0.03 0.157 1 7.04 0.000 1 2.55 0.000 3 0.35 0.000
Body part 3 0.069 0.015 .17 0.000 3 0.163 0.08 1 0.37 0.000
Body parts x Location 2 0.117 0.005 3 0.49 0.000 3 0.106 0.034 3 0.45 0.000
Error 11 0.013 -- 16 0.01 -- 15 0.028 -- 16 0.03 --
Parameters Body parts Locations
Vanadium Roe Spines Teeth Test Tripoli Saida
Nickel Spines Teeth Test Roe Tripoli Saida
Cadmium Spine Teeth Test Roe Tripoli Saida ______
Lead Roe Teeth Test Spines Tripoli Saida
Vanadium concentration was highest in the teeth and was lower in test, roe and spines (Figure 6 a). The level of nickel was still higher in spines than in other body parts (Figure 6 b). All 24
Chapter 2.Temporal variation
body parts in sea urchins from Tripoli had more cadmium in them than the sea urchins from Sidon (Figure 6, Table 3). The cadmium spines level in Tripoli had the highest value followed by teeth then test and roe (Figure 6 c). The mean level of lead was the highest in the teeth of sea urchins from Tripoli followed by test then roe with spines having the lowest mean of accumulated metal (Figure 6 d).
Figure 6 Metal bioaccumulation in µg of metal g-1(±SE) of sea urchin P. lividus’ body parts from two locations Sidon, South of Lebanon and Tripoli North of Lebanon for the year 2007 both at a distance of 600 m from a major pollution source (a-Vanadium, b-Nickel ,c-Cadmium and d-Lead).
Background metal levels in seawater and percentage body burden
Metals The mean concentrations of vanadium, nickel and lead (±SE) in sea water in Sidon were plotted against years. Each mean is the average of all distances (two points were taken in
25
Chapter 2.Temporal variation each of the furthest six sites equivalent to six points per year (Figure 7). Since sampling was carried out only once each year in May (water samples were taken 5 cm from the bottom of the sea floor), I included all metals results from the five years in the mean calculation of metal background concentrations in sea water (vs including only years 2006 and 2007 which may have resulted in the bioaccumulation of urchins in years 2007 and 2008) to minimize variability associated with snapshot sampling of the water body.
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Chapter 2.Temporal variation
Figure 7 Sidon‘s seawater background levels for a-vanadium, b-nickel, c- lead in µg l-1 (±SE) across years(2006-2009) averaging the 3 distances (450, 600 and 750 m). Results of One Factor ANOVA test are also shown.
The levels of each metal for each body part for all distances were summed and the percentage from the total sum of each metal was calculated and tabulated for each body part (Table 4). It is referred to as the percentage body part burden.
Table 4 Percentage body burden (dry weight) distribution for each of the four metals vanadium, nickel, cadmium and lead in different body parts of P.lividus in 2007.
Nickel Vanadium Cadmium Lead
% in body parts % in body parts % in body parts % in body parts
Roe 6.8 51.5 21 72.6
Test 24.8 20.6 34.2 12.6
Spines 59.7 3.1 25.6 2.6
Teeth 8.7 24.7 19.2 12.2
The mean water concentration of vanadium fluctuated from year to year with an average of 6.96 µg l-1 (Figure 7 a). This metal was accumulated mostly in the roe, while the teeth and test shared the rest of the percentage and the spines accumulated very little (Table 4).
The mean water concentration of nickel ranged, from 8.7 to 10.1 µg l-1, with an average of 9.65 µg l-1 (Figure 7b). This metal was accumulated mostly in spines followed by the test while teeth and roe accumulated very little (Table 4).
27
Chapter 2.Temporal variation
For cadmium, considering both its detection limit with the ICP-MS which is 0.2 µg l-1 and the fact that its bioaccumulation has been calculated in most body parts of P. lividus, it is safe to suggest that if cadmium existed in sea water, it must have been in quantities less than the detection limit. Therefore, the detection limit has been used as an upper range. The cadmium distribution percentage was very close in most body parts except in the test, which was higher by 1.5 fold (Table 4).
In seawater, the average of the lead measurement fluctuated from 0.12 to 0.29 µg l-1 (Figure 7 c). The roe had the highest percentage of lead while the spines accumulated the least (Table 4).
Chemical and Bacteriological
During the 5 years of census, sea water from Sidon location was analysed for bacteriological and chemical content in addition to metal concentrations (Table 5).
Table 5 Chemical and bacteriological characteristics of seawater (Mean ±SE) off Sidon harbour over the last five years (2005-2009) with the recommended safety levels of Marine Bodies for some parameters.
Parameter Mean (±SE) Recommended safety levels Reference for Marine bodies for various activities Total & fecal coliforms 30935 cfu ±17858 150 fecal coliform/100ml a ANZECC c -1 b Ammonia 22.6 mg l ±2.1 460 µg total NH3–N/L Batley,2009 Nitrates 2.52 mg l-1 ±0.7 -- -- pH 8.26 ±0.05 8.0-8.4 ANZECCd Conductivity 56.2 mS/cm ±0.49 -- -- Turbidity 100 NTU (taken @ 200 meter from pollution source) 0.5-10 NTU ANZECC d a Guidelines for recreational water quality and aesthetics b Trigger value for 95% protection concentration (Batley et al. 2009) c Recreational value (ANZECC 2000a) d Default trigger values for physical and chemical stressors for south-east Australia for slightly marine disturbed ecosystems
The purpose of comparing physical, chemical and biological parameters values in table 5 to international guidelines is to classify the disturbance level/status of Sidon‘s location in order to assist water resources managers in ensuring that aquatic ecosystems are adequately protected from disturbances. Most of the parameters listed in table 5 are considered ―Stressors directly toxic to biota‖ (ANZECC 2000a) . The level of total coliforms depicts a contamination level 200 times the maximal level recommended for recreational activity. In
28
Chapter 2.Temporal variation
Sidon sea water, the concentrations of ammonia that are protective of 99, 95 and 90% of species, as set by ANZECC, have been exceeded by 25 fold (2000b) ( 500-1220 µg total NH3–N/L) and by 50 fold the level set by Batley and Simpson (2009). These values are termed as trigger values by ANZECC, so that if the values are exceeded, further action must be triggered. The conductivity depicts a salinity of more than 40 ppt. Turbidity was measured only once at the 200 m distance from the pollution source and was greater than 100 units.
Discussion
The abundance of sea urchin Paracentrotus lividus has fluctuated over the five years of the study with two factors equally considered as possible causes of fluctuation (contamination and overfishing).The year 2007 scored the highest abundance and biomass of Paracentrotus lividus over the censused years in Sidon. But in 2008, even though the abundance was less, the metal content was higher than in 2007. In that year, P.lividus in Tripoli were more abundant and had higher biomass, gonad index and metal content than P.lividus in Sidon.
The scarcity of previous spatial and temporal variation studies for P. lividus in Lebanon and specifically Sidon, renders it difficult to compare the current situation with that of previous years. Local fishermen state that sea urchins were so abundant in the 1960s and 1970s that it was dangerous and quite difficult for swimmers to enter the water without foot protection. The density and the biomass of the sea urchin P. lividus in Sidon seawater are now relatively low compared to other fished/contaminated areas in the Mediterranean region (see summary of published literature in Table 6 where densities are ranked in the range reported for the Mediterranean Sea). Throughout this study, the size of P. lividus was within the normal size range for the Mediterranean basin although this is generally quite small in relation to other European locations. As far as distance from the major pollution source, the fact that the 300 m transects were void of urchins is possibly due to the contaminated condition of the location and to the waste deposit created. The physical and chemical parameters of the seawater, more generally outside Sidon harbour, reveal a contaminated area, which may be one of the major reasons for the depletion of P. lividus density besides overfishing. However, the very low densities of urchins in every year, except 2007 may relate to recreational fishing activity which was banned in the year preceeding 2007. The impact of overfishing is suspected in the absence of sea urchins at 750 m distance from pollution source which is close to a recreational fishing hotspot. Tripoli data reflects the importance of Marine Protected Areas in terms of relative high density and biomass of sea urchin P. lividus in a location partially sheltered from fishing. Of course it would have been better if more protected marine areas were sampled to be able to compare with fished areas but unfortunately the Tripoli location is the only marine protected area in Lebanon. Tripoli and
29
Chapter 2.Temporal variation
Sidon sea urchins bioaccumulated metals in their body parts, and this is possibly due to continuous industrial and domestic discharges, despite the fact that the bioconcentration factor was different in both locations, and different in Sidon in different years.
30
Chapter 2.Temporal variation
Table 6 Comparison table for density (number of individual m-2), size (test diameter in cm without spines) and biomass (g of wet weight m-2) of sea urchin P. lividus between Lebanon and other region of the Mediterranean. Locations are classified according to their status as reported by authors: Protected (P), F for fished, R for reference, and NM or the asterix (*) if not mentioned. Depth is expressed in m.
Country Location Status Depth Density Biomass (g Size (cm) Authors, Years (m) (m-2) wt wt m-2) Lebanon P 3-12 11.4 58.6 3.7 This study (Tripoli) Lebanon F 3-7 0.03-2 1.8-30 3.5 This study (Sidon) Croatia NM 0-3 * * 1-6.5 (Tomsic et al. 2010) Italy P 6-10 3-5.8 * 5.5-6.2 (Ceccherelli et al. 2009) Italy F 6-10 2-12 * 2.3-5 (Ceccherelli et al. 2009) Italy P 3-6 1.6-10.2 35-250 3-6 (Chiantore et al. 2008) Italy P 3-5 15.8-23 164.5 4.8-5.5 (Pais et al. 2007) Italy F 3-6 8.6-14 872-1555 3.2-3.3 (Pais et al. 2007) Turkey NM NM 31.6-59.4 4.1-4.8 (Dincer et al. 2007b)
Italy P 4-8 2.4-4 * 4.2-4.5 (Gianguzza et al. 2006) Italy F 4-8 0.8-3 * 3.5-4 (Gianguzza et al. 2006) Morocco NM NM * * <1-4.5 (Bayed et al. 2005) Italy P 5-6 11-15 125-280 2.8 (Guidetti 2004) Italy F 5-6 12 50 1.9 (Guidetti 2004) Spain P 2.5-10 4.7-15 * * (Tomas et al. 2004) Spain NM 3-10 2.1-7 * * (Sanchez-España et al. 2004) Italy P 8-12 7-12 130-160 2-3 (Guidetti et al. 2003a) Italy F 8-12 7-12 250-300 3-4 (Guidetti et al. 2003a) France NM NM * * 2-4.5 (Fernandez et al. 2000) France NM 1-8 0.8-5.6 * 3-5 (Ruitton et al. 2000) Tunisia NM 3 * * 2.4-6.2 (Sellem et al. 2000) France NM NM 0-6 * * Verlaque and Nedelac 1983a, Palacin et al.., 1998a Spain P 5-10 3.8-8 * <1-7 (Sala et al. 1998) Spain F 5-10 8-16 * <1-6.5 (Sala et al. 1998) Spain NM 13-16 3.4 * * (Palacin et al. 1998) Spain NM 13-16 0.9-3.4 * >3 (Palacin et al. 1998) Corsica NM 0.5-1 0.7 * * Bouderesque et al.., 1997 b Spain NM 4-16 2 * * Palacin et al.., 1997 b Spain P 5-10 2.4-5.1 * 2.3-3.9 (Sala et al. 1996) Spain F 5-10 8.4-10.7 * 3.1-3.6 (Sala et al. 1996) France NM NM * 650 * San Martin,1995 a Spain F* 0.5-0.7 3-31 * 1-5 (Turon et al. 1995) Spain P* 3-10 2-11 * 1-7 (Turon et al. 1995) France P 1-2 10 * 4.6 (Augier et al. 1994) France F 1-3 0-12 * 3.6-5 (Augier et al. 1994) Italy NM <1 24 * * Benedetti-Cecchi and Cinelli,1993 b France NM 2-5 0.7 * * Delmas 1992 b France NM NM 10-30 * * Fenaux et al.. 1987 a Portugal NM 0-2 <1 * * Saldanha, 1974 b Malta NM 5 4 * * Gamble, 1965 b Ireland NM 0.5 7 * * Kitching and Ebling, 1961 b a References extracted from Tomas et al., (2004) b References extracted from Boudouresque and Verlaque (2001) 31
Chapter 2.Temporal variation
Density, Biomass and Size Although settlement and abundance of sea urchins can vary greatly over time (Sala et al. 1998, Hereu et al. 2004), the variation I observed in the density of sea urchin P. lividus in Sidon‘s harbour may also be due to a combination of pollution and overfishing: I have observed lower fish stocks in areas that are heavily fished (pers.obs.). Similarly, when these fishing activities came to a complete halt in the aftermath of the 2006 war in South Lebanon (after a complete sea blockade was undertaken from July to November 2006), I noticed a partial recovery of the sea urchin stock (density increased from 0.033 to 1.55 individual m-2). This increase is perhaps due to the high reproductive potential of P. lividus, as they become sexually mature at a far smaller size in intensely exploited areas as indicated by Lozano et al.(1995) and Bouderesque and Verlaque (2001). Migration, due to the search for food (Crook et al. 2000, James 2000) could be another factor affecting the abundance of the sea urchin P. lividus , however most sea urchins display homing behaviour (James 2000, Tuya et al. 2004, Hereu 2005). Migration in this case is not considered a principle factor.
Predation is not likely to be a key determinant of density, as the main predator, the white sea bream (Diplodus sargus) (Sala et al. 1996) is overfished in the area. As with many other fisheries, its stock is depleted (pers.comm.). Other predators, such as sea stars and crabs for instance, were not observed during most of the undertaken dives. Few sponges and mussels (both fished) were spotted during dives but in very low counts (Figure 8). It is safe to say that biodiversity is not at its best in this area. It is also worth mentioning that in Sidon‘s harbour, very few sightings of Arbacia lixula were recorded from 2005 to 2009 (<5 per 1000 square meter). Arbacia lixula and Paracentrotus lividus are the two species of sea urchins dominating the coastal waters of Lebanon, however unlike P. lividus, A. lixula is not fished, as its roe is not palatable. When P. lividus is heavily fished along the Mediterranean, an increase in density of A. Lixula is noticed (elimination of competitor) (Guidetti 2004). However, despite the fact that A. lixula is not fished for commercial purposes in Lebanon (unprofessional recreational users might mistake it for the P.lividus and hence fish it) and even though P. lividus is heavily fished, the density of A. lixula remains the lowest in the area, which suggests that there may be other factors, besides fishing, behind this low density.
32
Chapter 2.Temporal variation
a
b
c
Figure 8 Sea urchin Paracenntrotus lividus and other biota off Sidon‘s harbour, 2007. a- a sea urchin (3.8cm), algae and a fish, b-sea urchin (4.2cm) on rocky substrate, c-sponges few meters away. Photos were taken on one of the dives in 2007 off Sidon‘s harbour during the fieldwork.
33
Chapter 2.Temporal variation
According to Sala and Zabala (1996), the bias selection by fisherman and predators of the sea urchin P.lividus affects its size range. For instance, predators such as fish and crabs usually target the 2-4 cm sea urchins as they are easy to crack, while fisherman tend to target bigger sea urchins >4 cm while the sea urchins < 2cm shelter in crevices to evade predation. The sea urchin size outside Sidon‘s harbour ranges from 3.31 to 3.83 cm, and is typical of Mediterranean Sea, indicating intense harvesting be it commercial or recreational rather than natural marine predators.
The contamination in Sidon‘s sea water (Table 5) can be split into 3 major categories: physical, bacteriological and chemical. Physical pollution such as the litter (mainly plastic and metallic containers) ejected from the waste deposit mountain (Zaatari 1998), which has already covered ca 30-40% of the rocky substrata in the neighbouring sites (200 to 350 m from pollution source) (Figure 9), is making it substantially difficult for the sea urchin to remain in its habitat. This littering of the sea floor may force P. lividus to migrate or it will possibly increase its mortality rate especially after settlement. This is mostly seen in the 300 m distance from the major source of pollution where sea urchins were not observed on the rocky substrata and algae of Sargassum species were dominant and formed forests covering huge areas in patches where litter was absent. This algae is known to survive in metal contaminated water where it bioabsorbs heavy metals (Roberts et al. 2008, Patrón-Prado et al. 2010).
The high turbidity (table 5) caused in general by the suspended solids stops the light from reaching the benthic communities sometimes even at low depths (5-7m). This is one of the reasons why the divers could not undertake any activities closer to the pollution source. Another physical pollution heavily weighing in South Lebanon, is that of blast fishing (Editor 2007, Zaatari 2007a), an illegal activity with damaging consequences to marine habitats.
In addition to the physical pollution, sea urchins in Sidon‘s harbour suffer from bacteriological contamination. In fact, sea urchin P. lividus accumulates Salmonella, Pseudomonas aeruginosa and many other pathogens (Portocali et al. 1996). The biological quality of the sea water off Sidon‘s harbour indicates the presence of high fecal and total coliforms. This presence signals the existence of other pathogenic viruses, which in turn will lead to an increased mortality rate. Therefore, mortality affects reproduction, hence density when the pathogens are bioaccumulated by sea urchins (Turon et al. 1995).
34
Chapter 2.Temporal variation
Figure 9 Litter photographed off Sidon‘s harbour seabed floor where cans, plastic and other material are covering the rocky substrate in one square meter and disturbing the sea urchin Paracentrotus lividus in its own habitat. In other places the litter was completely covering the sea floor bed.
For sea urchins collected from North Lebanon, the higher density may be attributed to the fact that the location is nested in the Marine Protected Area of Palm islands. This relatively high density in a protected area in Tripoli is in agreement with Gianguzza et al. (2006). The biomass of sea urchin in this protected area is higher than its southern counterpart which is in total agreement with other studies done on biomass (Guidetti 2004, Pais et al. 2007) and references therein where it was also noticed that the average of biomasses of P. lividus at the fished
35
Chapter 2.Temporal variation location was lower than the control. Biomass can increase in contaminated areas due sometimes to high organic content of domestic effluents (Welch et al. 1992).
Arafa (2006) and Bayed et al. (2005) both state that there is a direct relationship between food availability and gonad development. The values of the gonad index of the sea urchin P. lividus in Lebanon are above that reported by Tomsic et al. in Croatia (2010) (1.33 to 4.83) and more in agreement with the ones reported by Chiantore et al. (2008) (2.2-8.9) in field and by Spirlet et al. (2000) 9-12% in a laboratory experiment optimizing the gonad index through light, temperature and diet manipulation. The gonad index fluctuates according to the hydrodynamic conditions (Chiantore et al. 2008) and quality of food (Spirlet et al. 2000, Garmendia et al. 2010). This can probably explain the differences in gonad index and roe weight between Sidon and Tripoli where a smaller disturbance to light and food availability in the latter might favour the growth of roe and hence the gonad index.
Metals in Tissues The chemical pollution is a potential determining factor of density. Metals, released in the sea water of Sidon‘s harbour, were accumulated directly or indirectly in the body parts of the sea urchin P.lividus. Lead and cadmium, in particular, have been known to affect reproduction and hatching success rate of the sea urchin (Au et al. 2003) hence affecting their abundance.
The metals investigated were selectively distributed among the P. lividus body compartments. They did not differ with distance from the pollution source. This is perhaps due to the presence of other polluting sources around the waste deposit mountain contributing to the elevated levels of some of these metals. Alternative polluting sources are: port activities (a major source for nickel and vanadium), industrial discharges from the tanneries and soap factories around the harbour and the regular domestic waste discharged untreated into the sea water of Sidon‘ harbour.
In comparison with other locations in the Mediterranean Sea and with background levels from other international locations in the world (Table 7), I compared the mean levels for each metal for all years specifically against: 1) The Australian and New Zealand Environment Commission Council (ANZECC 2000b) background level, 2) the world background levels and 3) the Mediterranean Sea back ground level as well as the Mediterranean coastal level for each of the metals of interest (except for vanadium, which was compared to values taken from Fichet and Miramand (1998)). I did not find any record of the vanadium level in the Mediterranean sea to the best of my knowledge, however I found it to be from 1-3µg l-1in oceans with up to 6µg l-1 in coastal industrial areas (Fichet et al. 1998). The nickel level was measured as it was significantly higher than any of the references used for comparison. Lead level is in the range observed in the Mediterranean coast and its bioaccumulation in the consumed roe of P. lividus in Sidon has exceeded the food safety levels for lead and this should trigger an action from the 36
Chapter 2.Temporal variation authorities. As for cadmium, even though it was not detected in sea water by our instrument, however, its presence in the seawater can be expected at one period in time due to the fact that it was accumulated in all body parts of sea urchins from Sidon and Tripoli.
Table 5 Metal values of sea water (Mean µg .l-1 ±SE) off Sidon harbour over the last five years (2005- 2009) with the background levels in the Mediterranean and in other water bodies around the world listed for comparison
Parameter Mean µg L-1 Coastal Open Australia-New USA range for World range (±SE) Mediterranean Mediterranean Zealand range reference for reference sea range sea range for for reference background background sea water of -1 -1 -1 Sidon-Lebanon µg L reference background µg L µg L background µg L-1 µg L-1 1-3 µg L-1 (Ocean),6 µg Vanadium 6.9 µg L-1(±0.47) * * * * L-1 (coastal*)
Nickel 9.7 µg L-1 (±0.32) ND <0.5 0.13-0.5 0.3-5 0.12-0.7
Lead 0.2 µg L-1 (±0.03) 0.016-20.5 0.018-0.14 <0.006-0.03 0.01-1
Cadmium <0.2 µg L-1 0.002-0.9 0.004-0.06 0.025-0.38 0.01-0.2 0.001-1.1
(Manfra et al. (Manfra et al. (ANZECC (ANZECC (ANZECC References 2005) 2005) 2000b) 2000b) 2000b)
*No levels were available for vanadium from these sources; values for comparison were taken from Fichet and Miramand (1998). The relative size of the sea urchins collected in the year 2007 suggests that they were approximately 2-3 years old (Crapp et al. 1975). With a lack of a fully developed roe especially in the first few months during winter seasons and prior to reaching sexual maturity, most of the accumulation would have happened in the test, teeth and spines of the urchin as seen in Warnau et al. (1998). As roe begins to grow in size leading to spawning, accumulation increases in the roe (because more binding sites become available) before spawning occurs and toxicants get purged. This increase in levels of vanadium and lead in roe of Sidon‘s P. lividus is probably linked to an increase in lead and vanadium levels in seawater at the time of roe maturation which could not have been detected by one single yearly seawater sample.
In 2008, the metal concentrations in the body parts of sea urchins in Sidon depicted levels much higher than the metal concentrations in the body parts of sea urchins in Sidon in 2007. This may be due to two reasons: i) sea urchins in Sidon 2008 have been accumulating metals for at least two years (2007-2008) whereas the sea urchins collected on 2007 have been accumulating the metals of year 2007 only since I did not find much of urchins in 2006, ii) because following the complete destruction of the power plant after the war activities of 2006, all factories in the south of Lebanon feeding from this plant came to a standstill from July 2006 to May 2007 which probably decreased the discharge rate of heavy metals into the surrounding environment, iii) the
37
Chapter 2.Temporal variation massive exodus of the people of the South from July 2006 to March 2007 which decreased the domestic waste discharged into the area.
As for Tripoli, North of Lebanon, despite the fact it is a marine protected area as far as fishing activity is concerned, unfortunately many pollution sources discharge into the area e.g.: the Port of Tripoli with all the boating activities, domestic and industrial wastes from local homes and factories respectively. Since sea urchins from this location are left ―undisturbed‖ from fishing, they have presumably been accumulating metals over the years exceeding by this the levels of the metals measured in sea urchins of Sidon during the same year. However the metals‘ levels are still within the overall range measured in the Mediterranean basin. In general, I noticed that the results of metal accumulation in the body parts of the sea urchins P. lividus are well within the established ranges seen in the Mediterranean Sea as listed in Table 8. The emphasis in most of the papers listed was on the roe since it is the edible part. Unfortunately few papers listed the spines, test and teeth bioaccumulation values. The lead concentrations in Tripoli were higher in calcites than in roe. This has also been noted by Temara et al.(1997), who noted a similar thing in the calcites of sea star A. rubens in low background concentration environment. The size of the sea urchin has not shown a relationship with increasing or decreasing metal bioaccumulation in any of the body parts (Table 8). It is worth noting that sea urchins sometimes invest in reproduction at the expense of the somatic growth (Bayed et al. 2005). This implies that sea urchins of similar sizes might be of different ages. Recent studies, unfortunately, have shown that the climate change can also negatively impact the growth and the reproduction of the sea urchin, which predict a dim future for the sea urchin population (Byrne et al. 2010).
38
Chapter 2.Temporal variation
Table 8 Comparison table for bioaccumulation of cadmium, lead and nickel in different body compartments of the sea urchin P. lividus in Lebanon with other region of the Mediterrenean. Locations are classified according to their pollution status as reported by authors: Contaminated (C, even if locations/sites are lighlty contaminated), R for reference, and NM if not mentioned . Depth is expressed in m, Analytical technique used for the final determination of metal bioaccumulation: GF-AAS (Graphite Furnace-Atomic Absorbtion Spectrometery), AES (Atomic Emission Spectrometry), ICP-MS (Ion Coupled Plasma- Mass Spectrometry),N number of sea urchins used for analysis of metals, their mean and/or range size in cm, dry or wet samples used in bioaccumulation calculations, mean and/or range of metal values (all in dry weight) pooled from sites within locations, authors and year.
Metal Year of Country Location Depth Analytical N Size cm Roe Test Spine Teeth Reference study status m Technique µg Kg-1 µg Kg-1 µg Kg-1 µg Kg-1
Cadmium 2007-2008 Lebanon C 3-12 ICP-MS 93 3.7 <20-117 191-4535 143-1878 107-3815 This work Cadmium 2007 Lebanon C 3-7 ICP-MS 12 3.5 371 424 2766 600 This work Cadmium 1991-2005 Spain C NM GF-AAS NM 4.3 140-1590 NT NT NT (Deudero et al. 2007) Cadmium 2002 Algiers C 1-5 GF-AAS 10 4.5-6.5 50-140 NT NT NT (Soualili et al. 2008) Cadmium 2002 Algiers C 1-6 GF-AAS 10 4.5-6.5 50-140 NT NT NT (Soualili et al. 2008) Cadmium 1999-2000 Morocco C NM AES 10 3.0-3.5 1510-25150 NT NT NT (Bayed et al. 2005) Cadmium 1998 Italy NM NM GF-AAS 1500 9 100-650 NT NT NT (Storelli et al. 2001) Cadmium 1991-1992 Corsica R 5-10 AES 7 6.3 240 175 NT 227 (Warnau et al. 1998) Cadmium 1991-1992 Italy C 5-10 AES 9 5.1 480 177 NT 252 (Warnau et al. 1998) Cadmium 1991-1992 France C 5-10 AES 9 6.4 452 175 NT 262 (Warnau et al. 1998) Cadmium 1991-Dec Italy C NM AES 74 1.2-6.5 NT 90-800 NT NT (Warnau et al. 1995a) Cadmium 1992-May Italy C NM AES 45 1.2-5 NT 80-410 NT NT (Warnau et al. 1995a) Cadmium 1992-1993 Egypt C NM AAS 15 NM 250-480 NT NT 2160-4390 (Mostafa et al. 1995) Cadmium 1993 Ireland NM NM AAS 16 NM 330-1020 NT NT 5430-7230 (Mostafa et al. 1995) Cadmium 1992 France R 1-2 AAS 10 4.6 <300 <300 <300 <300 (Augier et al. 1994) Cadmium 1992 France C 1-2 AAS 10 3.6-4.6 <300 <300-2300 <300-1600 <300 (Augier et al. 1994) Cadmium 1988 Greece R 12 AAS NM NM 1900-3400 NT NT NT (Castiki et al. 1991) Cadmium 1979 Lebanon C 1.5-2 AAS 12 3.8 400-4900 NT NT NT (Shiber 1979) Nickel 2005-2009 Lebanon C 3-12 ICP-MS 93 3.7 227-2924 833-11895 2038-10971 300-7475 This work Nickel 2007 Lebanon C 3-7 ICP-MS 12 3.5 1096 150 206 290 This work Nickel 1991-2005 Spain C NM GF-AAS NM 4.3 800-2700 NT NT NT (Deudero et al. 2007) Nickel 1992-1993 Greece C NM AAS 30 NM 4640 NT NT NT (Portocali et al. 1996)
39
Chapter 2.Temporal variation
Table 8 Continued
Metal Year of Country Location Depth Analytical N Size Roe Test Spine Teeth Reference study status m Technique cm µg Kg-1 µg Kg-1 µg Kg-1 µg Kg-1
Nickel 1992-1993 Greece R NM AAS 34 NM 7200 NT NT NT (Portocali et al. 1996) Nickel 1992-1993 Egypt C NM AAS 15 NM 2130-9420 NT NT 11050-21960 (Mostafa et al. 1995) Nickel 1993 Ireland NM NM AAS 16 NM 590-1270 NT NT 28890-39850 (Mostafa et al. 1995) Nickel 1988 Greece R 12 AAS NM NM 7100-34900 NT NT NT (Castiki et al. 1991) Nickel 1979 Lebanon C 1.5-2 AAS 12 3.8 <10-40300 NT NT NT (Shiber 1979) Lead 2005-2009 Lebanon C 3-12 ICP-MS 93 3.7 198-2924 37-743 7-10971 35-7475 This work Lead 2007 Lebanon C 3-7 ICP-MS 12 3.5 309 477 340 654 This work Lead 1991-2005 Spain C NM GF-AAS NM 4.3 800-6700 NT NT NT (Deudero et al. 2007) Lead 2002 Algiers C 1-5 GF-AAS 10 4.5-6.5 800 NT NT NT (Soualili et al. 2008) Lead 2002 Algiers C 1-6 GF-AAS 10 4.5-6.5 880-7780 NT NT NT (Soualili et al. 2008) Lead 1999-2000 Morocco C NM AES 10 3.0-3.5 5220-35320 NT NT NT (Bayed et al. 2005) Lead 1998 Italy NM NM GF-AAS 1500 9 100-2650 NT NT NT (Storelli et al. 2001) Lead 1997 Spain C NM ICP 10 NM NT 60000 21000 NT (Chinchon et al. 2000) Lead 1991-1992 Corsica R 5-10 AES 7 6.3 1380 2520 NT 2192 (Warnau et al. 1998) Lead 1991-1992 Italy C 5-10 AES 9 5.1 1930 2577 NT 2155 (Warnau et al. 1998) Lead 1991-1992 France C 5-10 AES 9 6.4 1660 2882 NT 2275 (Warnau et al. 1998) Lead 1994 Spain C 1-2 ICP 10 NM NT 59800 21000 NT (Auernheimer et al. 1997) Lead 1994 Spain R 1-2 ICP 10 NM NT 2200 2200 NT (Auernheimer et al. 1997) Lead 1991-Dec Italy C NM AES 74 1.2-6.5 NT 210-370 NT NT (Warnau et al. 1995a) Lead 1992-May Italy C NM AES 45 1.2-5 NT 270-420 NT NT (Warnau et al. 1995a) Lead 1992-1993 Egypt C NM AAS 15 NM 840-2990 NT NT 10640-22230 (Mostafa et al. 1995) Lead 1993 Ireland NM NM AAS 16 NM 630-1200 NT NT 34150-45110 (Mostafa et al. 1995) Lead 1992 France R 1-2 AAS 10 4.6 800 <500 <500 <500 (Augier et al. 1994) Lead 1992 France C 1-2 AAS 10 3.6-4.6 <500-3300 <500 <500 500 (Augier et al. 1994) Lead 1979 Lebanon C 1.5-2 AAS 12 3.8 11300-313000 NT NT NT (Shiber 1979)
40
Chapter 2.Temporal variation
The lead level measured in the roe of the sea urchin P. lividus (2.92 ppm) from Lebanon (Sidon 2008 sampling) is much higher than the safety range (0.3- 1.5 ppm) set by the FAO/WHO (2009) , FDA (2001) and EU (2006) collectively. Although some of the values in the legislation have been set up for Mollusc and Cephalopods, it has been customary for scientists to apply this limit for other sea food products (Storelli et al. 2001). As for the cadmium and nickel levels in the sea urchin P. lividus, these were found to be below the safety limit. To this date, no recommended safety limit has been established for Vanadium in food.
Conclusion
In general, it can be concluded that Sidon‘s harbour is a contaminated location with a major source of pollution affecting the sites in close proximity to it, though it is not clear if the waste deposit mountain is the sole source of contamination relative to other potential sources. However, this contamination cannot be deemed individually accountable for the fluctuation of density across the years and across the distances from main polluting source, for this has been achieved in conjunction with overfishing. The distance from the pollution source did not affect the levels of tissue contamination with metals, which were not equally distributed in the different body parts of the sea urchin Paracentrotus lividus. The values of measured metals in different body parts are quite close to those measured in other polluted parts of the Mediterranean, even though industrial activities in Lebanon, regardless of scope, remain much smaller than the rest of the Mediterranean countries reviewed in this study. The lead levels in the roe of the sea urchin are higher than the international recommended levels, which is in itself a valid reason for the authorities to push for stricter restriction on fishing and to biomonitor the metals in the marine biota. Depuration experiments must be undertaken to assess ability of P. lividus to purge toxicants like lead and cadmium before declaring this organism safe for human consumption.
This study may act as a baseline for future biomonitoring and subsequent studies. It is recommended that more in depth studies be taken in bioaccumulation, especially in metals for which there is a paucity of papers (such as vanadium) to better understand the differences in partitioning of metals in body parts of the sea urchin P. lividus. However, as far as Sidon‘s harbour is concerned, it is recommended that a management program be implemented to reduce the discharge of toxic substances. It should include monitoring the abundance (for which no conclusion can be made regarding its temporal stability) and distribution of contaminants in the Sidon Bay from season to season in order to determine the effectiveness of control efforts. Also parameters such as salinity, turbidity, temperature and food availability must be monitored more often as they affect the gonad index of the P. lividus sea urchin,
41
Chapter 2 Temporal variation
where the results suggest that it can be used as an effective biomonitor for this location. With regard to fishing activities, , more stringent measures should be taken, which could include regulating quota and size of fished P. lividus and issuing fishing permits for professional fishermen.
42
Chapter 3
3.Bioaccumulation and Bioconcentration factors for a cocktail of inorganic and organic toxicants in the sea urchin Paracentrotus lividus
Abstract
To better understand the bioaccumulation of contaminants in the body parts of P. lividus, laboratory exposures to a cocktail of toxicants made of organics (PAHs, OCPs), inorganic metals (Ni,V, Cd, Pb) and bacteria were undertaken in different concentrations. The experiments followed the progress of toxicant bioaccumulation in the different body parts of the organism and determined their bioconcentration factor. In the roe, vanadium, nickel, cadmium and lead accumulation was linear after 28 days of exposure. The roe accumulated the most among all tested body parts. There was a two-week delay in the accumulation of metals in the calcites (test, teeth and spines) with increases in accumulation observed during the third and fourth weeks. The magnitude of accumulation was lowest in the teeth. Nickel accumulation was greater in the spines. The bioconcentration factors (BCF) of cadmium at the end of the 4th week ranged from 30 in the test to 400 in the roe. As for lead, the BCF ranged from 39 in the teeth to703 in the roe. Accumulation of OCPs (4,4‘DDT and dieldrin) were measured in the roe of the sea urchin along with phenanthrene, anthracene and pyrene . For the OCPs, after the second week BCFs were ranging from 7,348 for dieldrin to 36,604 for 4,4‘DDE (metabolised from 4,4‘DDT). BCFs of PAHs, after the second week, were 894,19,362 for phenanthrene, anthracene and pyrene respectively. P lividus is a good bioaccumulator of wide classes of toxicants, which has implications for fisheries management. The use of a cocktail of toxicants in this study is aimed to simulate a realistic contamination event that would result in potential antagonistic, additive or synergistic effects between toxicants. Future research should explore each of these toxicants independently.
43 Chapter 3.Bioaccumulation
Introduction
Bioaccumulation & Bioconcentration factor Terrestrial and marine environments are being modified by anthropogenic activity at a very high rate (Yokota et al. 2010). The assessment of ecosystem health involves the monitoring of biological indicators from molecular to community level in order to evaluate the hazards of contaminants. The assessment requires early identification of stress in biota before the uptake of toxicants has exceeded the lethal body burden, causing irreversible damage. International guidelines provide maximum permissible concentrations for some chemical, physical and microbiological entities which safeguard a certain percentage of biota. However, these guidelines for concentrations are subject to change as more and more studies are completed and outcomes of these studies lead regulators toward setting new guidelines with more stringent values for safeguarding human and animal health (ANZECC 2000a, Russo 2002, Matthiessen et al. 2007).
One of the tools used in ecotoxicological studies are laboratory exposure experiments, where bioconcentration factors (BCF) and persistent and acute toxicity of contaminants are determined (Boudou et al. 1997, McGeer et al. 2003). Bioconcentration refers to the process of accumulation of chemicals in an aquatic organism as a result of exposure of the organism to a chemical concentration in the water via non dietary routes (ATSDR 2008). The BCF is a measure of the tendency of a substance to bioconcentrate in aquatic organisms and may help to determine the toxicity of that compound/substance(Kumar et al. 2009) . BCF is also a fundamental parameter employed in environmental management strategies, such as the European Community Regulation (No 1907/2006), on chemicals and their safe use (Lombardo et al. 2010). In summary, BCF is a useful indicator of biological effects and has been employed to predict the hazards of toxicants to aquatic organisms.
Knowledge of bioaccumulation rates in many marine organisms is still not well understood. While there are dangers in extrapolating results from laboratory based experiments to the field, laboratory based exposure experiments have an important role to play in estimating the potential risk of contaminants (Lam et al. 2001) when they are discharged as waste into the aquatic environment .
Common marine contaminants The categorization of different wastes being discharged into the aquatic environment from the world‘s most active coasts has revealed a mixture of microbes, inorganic and organic compounds. The presence of these compounds, at different concentrations, have been
44 Chapter 3.Bioaccumulation observed in aquatic ecosystems worldwide (Tilghman Hall et al. 1991, Portocali et al. 1996, Tsigouri et al. 2000, Szlinder-Richert et al. 2008, Taniguchi et al. 2009).
Toxicants are taken up by organisms either directly through water or through diet. At the cellular level, metals can be taken up and fractioned into a metabolically active fraction and a detoxified fraction (inactive metabolically). The metabolically active fraction includes organelles (mitochondria, endoplasmic reticulum), heat resistant proteins (metallothionein in fish and mammals) and heat sensitive proteins (enzymes) and the detoxified fraction includes the metal rich granules and metallothionein-like proteins (found in invertebrates)(Ferrarello et al. 2000, Wang et al. 2006, 2008). The onset of toxic effects depends only on the concentration of accumulated metal in its metabolically available form (Rainbow 2002). For organic compounds, such as Polycyclic Aromatic Hydrocarbons ( PAHs) and Organo- Chlorinated Pesticides (OCPs), these are taken up at the cellular level by some membrane bound enzymes for example the Cytochrome P450 isoenzymes. Once the organic toxicants bind to redox metabolising enzymes, including the cytochrome P450 isoenzymes, metabolites are formed by oxidisation and synthesized into water soluble compounds which may or may not be easy to eliminate. Metabolites can be more toxic than the precursor compound, such as carcinogens, (Newman et al. 2008) and will be damaging to the cell and the corresponding tissues.
Furthermore, accumulated contaminants can have an impact on the immune system. Some of the metals, for instance, might have synergistic effects with other contaminants (Xu et al. 2010) creating even more complex physiological situations at the cellular level (Sharara et al. 1998, Pulido et al. 2003, Deane et al. 2006, Birceanu et al. 2008). For example, metals which occur as co-contaminants with organic xenobiotics (e.g. dioxins and PAHs) can increase reactive oxygen species formation through oxidative effects (direct and indirect mechanisms) and thus can potentially reduce P450 biotransformation efficiency of dioxins and PAHs (Vakharia et al. 2001, Benedetti et al. 2009) hence increasing sensitivity of the cell.
Inorganic contaminants in industrial effluents include many naturally occurring metals such as cadmium, lead, nickel and vanadium, among others. Toxic metals in particular have become a ubiquitous problem worldwide. Lead, cadmium and nickel are also products of mining activities. For instance, lead and cadmium, which have been detected in most water bodies (Manfra et al. 2005), can be toxic to living organisms , including humans. They may affect nervous and reproductive systems even at low concentrations, with sufficient evidence for carcinogenicity in humans (ATSDR 2007, 2008). The harmful health effects of nickel in humans range from allergic reactions such as skin rashes and asthma attacks, to more serious
45 Chapter 3.Bioaccumulation health effects such as chronic bronchitis, reduced lung function and cancer of the lung and nasal sinus (high exposure) (ATSDR 2005). Vanadium released into the environment are mainly associated with industrial sources, especially oil refineries and power plants which use vanadium rich fuel oil and coal (ATSDR 2009). As for its health effects, the International Agency for Research on Cancer (IARC) has determined that vanadium can be a potential carcinogenic to humans.
Organic pollutants include classes such as PAHs, OCPs, phenols, nitroaromatics and benzene derivatives. These compounds have a great affinity for lipid-rich tissues, are chemically stable, poorly metabolised (Moore et al. 2002) and can accumulate in various organs leading to toxicity in both humans and animals. This has prompted governments to regulate them and researchers to be concerned with their presence in the environment. For instance, OCPs are listed as ―Persistent Organic Pollutants‖ (POPs) in the Stockholm Convention (UNEP 2001). This convention requires in its Article 11.1.b to monitor POPs, particularly their presence, levels and trends in humans and in the environment.
Sources of organic contamination in marine systems vary widely. PAHs and other aromatic petroleum hydrocarbons contamination in marine environments stem from activities such as transportation and accidental spillage of oil, incomplete combustion of fossil fuels, other industrial processes as well as natural occurrences such as forest fires. These molecules are pollutants of concern because of their toxic and carcinogenic potential (ATSDR 1995). Chlorinated pesticides such as the dichloro-diphenyl-trichloroethane (DDT) and its metabolites dichloro-diphenyl-dichloroethylene (DDE) and dichloro-diphenyl-dichloroethane (DDD) along with dieldrin are only two of a large number of related compounds used for pest control and while currently prohibited in many countries, their illegal use continued for many years after the ban started in 1972 (Yang et al. 2006).
Other organic compounds such as benzene and its derivatives damage the bone marrow in humans and depress the immune system, although there is no evidence of biomagnification in the marine food chain (ATSDR, 2007). As for phenols, they are mainly germicidal and are used in formulating disinfectants. However, some contamination may be of concern since some of these compounds have mitochondrial disrupting properties (Tjeerdema et al. 1994, TenBrook et al. 2003).
The microbial contamination in marine system loads includes coliforms whose presence may indicate a higher risk of pathogens being present in the water (WHO 2003) .Untreated organic matter, containing fecal coliforms and discharged into aquatic bodies, can also be harmful to
46 Chapter 3.Bioaccumulation the environment if aerobic decomposition of this material reduces dissolved oxygen levels down to concentrations that are low enough to kill fish and other aquatic life (Libes et al. 2003). Bioaccumulation studies of microbial loads in a laboratory can be efficient in animals known to accumulate all categories of toxicants. When placed in a metal loaded medium, bacteria interact with some of these metals and might change their bioavailability either by breaking their ligands with other molecules, or, by taking them as phagocytes ( Panwichian et al. 2010). This can ensure that metals in media loaded with bacteria do not interact directly with binding sites inside the roe, and bacteria are instead taken up by coelomocytes of sea urchin. Coelomocytes neutralise bacteria and any other toxicants sequestered inside the bacteria (Chia et al. 1996). Adding bacteria to a mixture of toxicants in exposure experiments not only mimic real-life situations but also accounts for any synergistic, antagonistic or additive effects between bacteria and other toxicants.
The study organism Benthic animals are good candidates for the study of bioaccumulation of a wide range of contaminants and can reflect conditions occurring over time. Those with restricted mobility may be especially sensitive to pollution and other types of habitat degradation because they reside on bottom sediments where chemical contaminants accumulate (Thompsons et al. 1989, Reed et al. 2010). Echinoderms are a group of marine benthic animals which include sea stars, sea urchins and sea cucumbers. They are genetically more related to chordates than any other major invertebrate group (Sea Urchin Genome Sequencing Consortium,et al. 2006). About 70 % of sea urchin genes are shared with humans. Echinoderms may possess control mechanisms of physiological processes, in terms of molecules and actions, rather similar to those of vertebrates (Sugni et al. 2007).
Sea urchins are echinoderms with many environmental qualities that make them potentially useful as bioindicator species. They accumulate most kinds of contaminants, are quite sedentary which means they can give a better time integrated picture of local toxicants (Cebrian et al. 2003). The Sea urchin Paracentrotus lividus, in particular, inhabits the Mediterranean basin and is heavily targeted for consumption due to the delicacy of its roe (Sugni et al. 2007). This species has been studied at several levels from molecular to community and has been extensively used in LC50, larval toxicity and fertilisation experiments. This has helped scientists to study the fertilisation process and to use it In Vitro Fertilization (IVF) for humans (Gordon 2003).
With all this history however, only four studies have examined bioaccumulation in this specific urchin and none to my knowledge have been done to quantify PAHs and OCPs
47 Chapter 3.Bioaccumulation accumulation in P. lividus. Miramand et al. (1982) looked at the vanadium accumulation in body parts of the P.lividus using a radiotracer. They found that uptake of vanadium in whole organism was slow after 3 weeks and that it mostly favoured the calcites. Warnau et al. examined bioaccumulation of cadmium in the different body parts of P. lividus using low concentrations, and established that sea urchins mostly accumulated cadmium from water rather than food (Warnau et al. 1995b). Danis et al. examined the bioaccumulation of PCBs using radio labelled tracers. They found that the echinoids accumulate PCB153 more efficiently when exposed to them in water than when they were exposed to them through their food. The target body compartments of PCB153 were found to be body wall and spines when individuals were exposed via water, and gut when they were exposed via food. In summary, the urchin is an important organism in the near-shore ecology of the Mediterranean and sufficient data is available about its physiology to suggest that it may make a useful bioindicator species. Moreover, the urchin is commonly consumed and contaminant loads are therefore relevant to human health issues.
Aims I therefore addressed the following three aims: a) To examine the bioaccumulation of a mixture of toxicants in the sea urchin P. lividus collected from the Lebanese coast. b) To test for differences in bioaccumulation of toxicants in body parts of P. lividus over a period of four weeks. c) To determine a bioconcentration factor for each of the toxicants in the tested body parts.
Materials and Methods
Collection of Specimen (Sea Urchins) Sea urchins, all of the species Paracentrotus lividus, were collected on the 25th of May 2006 (during spawning season) by SCUBA diving between 5 and 10 m depth off the coast of Saadiyat (South of Beirut, Lebanon). Prior to experimentation, the sea urchins (190 in total, average size ca 4 cm± 0.1) were acclimatized to laboratory conditions for one week in nine glass aquaria tanks (60L) filled with natural sea water constantly aerated (Salinity: 38-41%o, temperature: 22± 1 °C, 12/12 h dark/light cycle). During the acclimatization period, the tanks were monitored daily and dead specimens were removed. The sea urchins were distributed evenly between the tanks (20 sea urchins/tank). During acclimatization, the sea urchins were fed only once at the end of the week and also before the change of water for the rest of the experiment.
48 Chapter 3.Bioaccumulation
After acclimatization, the water in the tanks was replaced with fresh sea water. A sea water subsample (100 ml) was taken from each of the nine tanks and made into a composite control sample prior to spiking the water. One sea urchin was randomly taken out of each of the nine tanks before the addition of the toxicants to determine the bioaccumulation baseline (sea urchins from this collection were labelled week 0). These sea urchins were placed in the refrigerator until processed on the following day.
Experimental Set-up and Spiking (Exposure)
Trial 1 The nine tanks were split for use in three treatments (3 replicate tanks per treatment) as follows:
The full treatment consisted of three 60 litre tanks spiked each with a 50 ml mixture of deionized water: acetone (50:50) containing a mixture of toxicants and achieving a final nominal tank concentrations of 5 µg/l for PAH compounds (pyrene and acenaphthene,), 10 µg/l for each of the cadmium, lead, phenol compounds and few OCPs and 25µg/l (25 ppb) for dieldrin, Endrin and 4,4‘DDT (Toxicants listed in table 1). In addition, 2 ml of a microbiological culture containing 730 cfu/ml (colony forming unit) of total coliforms and E.coli strains were added to the mixture at the last minute. This culture was isolated from seawater samples collected off the Lebanese coast, maintained in specialised fridges and identified using ISO 4832:2006(E) method (with modification) at a CAP (College of American Pathologists) accredited laboratory. This treatment is referred to in the results as ―The full treatment‖.
The ―half treatment‖ consisted of three 60 litre-tanks spiked each with a 50 ml mixture of deionized water: acetone (50:50) containing half the concentrations of the cocktail of toxicants used for the full treatment. As for the control treatment, it consisted of a 50 ml mixture of deionized water: acetone (50:50) free of toxicants and microbial culture.
Spiking preparations Organic toxicants (six chlorinated pesticides, two PAHs, six phenols, two petroleum derivatives of high purity purchased from purchased from Absolute Standards and Ultra Scientific listed in table 1), were diluted in their proper dissolving solvents before being added to the carrier solvent (acetone, Merck). Inorganic metal toxicants (cadmium and lead purchased from Agilent and Absolute Standard) were diluted in deionized water (acidified with HCl) before being added to the acetone fraction. The microbiological culture was added after the mixing of the organic and inorganic toxicants and right before the spike of tanks.
49 Chapter 3.Bioaccumulation
This cocktail of toxicants was an attempt to simulate an environment as close as possible to real-life environmental conditions. Not all toxicants added to the sea water tanks were meant to be tested in the sea urchin. Their presence was simply to contribute to the complexity of the chemical environment as it occurs in many places around the Mediterranean coast. Only dieldrin, 4,4‘ DDT/DDE, acenaphthene, pyrene, anthracene, phenanthrene, cadmium, lead and E.coli, previously observed as accumulating in marine mammals (Yogui et al. 2003, Tuerk et al. 2005), were tested in the sea urchin P.lividus.
The levels of chosen toxicants reflected some of the current concentrations around the Mediterranean (the natural habitat of sea urchin) especially the eastern coast and in water or sediments around the world (Tjeerdema et al. 1994, Mai et al. 2002, Kim et al. 2006, Kljakovic´-Gašpic´ et al. 2010, Li et al. 2010) with inflation of the OCPs levels to compress the test period into a relatively short and easily manageable duration.
50 Chapter 3.Bioaccumulation
Table 1. List of the toxicants added to the experimental tanks for both trials. Also listed is the full treatment concentration nominal values in µg l-1, the Chemical Abstract Service number (CAS), the molecular formula, the molecular weight (MW), the log of octanol to water partition coefficient (Kow) and the water solubility for each of the toxicants. Source: ATSDR.-USA
Category Classes Analytes Full treatment CAS M.Formula MW K ow Water solubility
Trial 1
-1 Organic OCP 4,4‘DDT 25 µg l 789- 02-6 C14H 9Cl5 354.49 6.91 0.025 mg/L @25oC -1 Dieldrin 25 µg l 60-57-1 C12H8Cl6O 380.91 6.2 0.11 mg/L
-1 Aldrin 25 µg l 309-00-2 C12H8Cl6 364.88 6.5 0.011 mg/L
-1 Endrin 10 µg l 72-20-8 C12H8Cl6O 380.91 5.6 200 µg/L
-1 Gamma-BHC 10 µg l 58-89-9 C6H6Cl6 290.83 3.72 17 mg/L
-1 Heptachlor 10 µg l 76-44-8 C10H5Cl7 373.29 6.1 0.05 mg/L
-1 PAH Acenaphthene 5 µg l 83-32-9 C10H6(CH2)2 154.2 4.08 1.93 mg/L
-1 Pyrene 5 µg l 129-00-00 C16H10 202.3 5.07 0.077 mg/L
Phenols 4-Chloro-3-methylphenol 10 µg l 59-50-7 C7H2ClO 141.58 --- Slightly soluble
2,4-Dichlorophenol 10 µg l 120-83-2 C6H4Cl2O 163 3.11 4500 mg/L
2-Nitrophenol 10 µg l 88-75-5 C6H5NO3 139.11 1.79 1400 mg/L @25oC
Phenol 10 µg l 108-95-2 C6H5OH 94.11 1.5 82.2 g/L
Pentachlorophenol 10 µg l 87-86-5 C6Cl5OH 266.34 5.01 14 mg/L @20oC
2,4,6-trichlorophenol 10 µg l 88-06-2 C6H3Cl3O 197.45 3.69 434 mg/L
-1 Petroleum 1,4, dichlorobenzene 5 µg l 106-46-7 C6H4Cl2 147.01 3.52 79 mg/L @ 25oC
-1 1,2,4-trichlorobenzene 5 µg l 120-82-1 C6H3Cl3 181.45 4.02 <0.1%
Others 2,4-dinitrotoluene 5 µg l 121-14-2 C7H6N2O4 182.14 2 270 mg/1000 ml
N-Nitrosodi-n-propylamine 5 µg l 621-64-7 C6H14N2O 130.19 1.36 9894 mg/L @25oC
51 Chapter 3.Bioaccumulation
Category Classes Analytes Full treatment CAS M.Formula MW K ow Water solubility
Inorganic Metals Cadmium 10 µg l 10108-64-2 CdCl2 5/2H2O 228.35 --- Freely Soluble
Lead 10 µg l 7758-95-4 PbCl2 278.11 --- 1.1 g in 100g
Biota Microbiological Total Coliforms 730 cfu/ml ------
E. coli 730 cfu/ml ------
Table 1 Continued
Category Classes Analytes Treatment CAS M.Formula MW K ow Water solubility
Trial 2
-1 Organics PAH Anthracene 3 mg l 120- 12-7 C14H 10 178.23 4.53 0.076 mg/L
-1 Phenanthrene 1 mg l 85-01-8 C14H10 178.23 4.53 1.2 mg/L
-1 Naphthalene 1 mg l 91-20-3 C10H8 128.2 3.34 Insoluble
-1 Pyrene 1 mg l 129-00-00 C16H10 202.3 5.07 0.077 mg/L
-1 Petroleum 1,4-dichlorobenzene 1 mg l 106-46-7 C6H4Cl2 147.1 3.52 79 mg/L @ 25oC
-1 1-chloro-2,4-dinitrobenzene 1 mg l 97-007-xx C6H3(NO2)2Cl 202.35 2.17 Insoluble
-1 Phenols 2,4-dichlorophenol 1 mg l 120-83-2 C6H4Cl2O 163 3.11 Slightly soluble
-1 Phenol 1 mg l 108-95-2 C6H5OH 94.11 1.5 82.8 g/L
-1 Inorganics Metals Cadmium 1 mg l 10108-64-2 CdCl25/2 H2O 228.35 --- Soluble
-1 Nickel 1 mg l 13138-45-9 Ni(NO3)2.6H2O 290.83 --- 238.5 g/100cc
-1 Vanadium 1 mg l 1314-62-1 V2O5 181.88 --- 0.1-1%
-1 Lead 1 mg l 7758-95-4 PbCl2 278.11 --- 1.1 g in 100 g
52 Chapter 3.Bioaccumulation
Immediately after the tanks were treated with the contaminants, the contents of each of the tanks were mixed by slightly increasing the air bubbling. Composite water samples (100ml) were then taken from each tank after one hour and pooled for each treatment to test the final concentration of the toxicants for metals (Table 7). Throughout the experiment, water was changed weekly, and one day prior to each water change, the sea urchins were fed algae (Ulva species) at around 2% of their body weights. The feeding schedule was organized as such to minimize the adsorption of toxicants to the food, which could lead to lower levels of free toxicants in the tanks. After each water change, the tanks were re-spiked with the toxicants as described in the process above. The tanks were monitored every other day for parameters such as oxygen, salinity, temperature and pH. Faeces and shed spines were removed daily to avoid adsorption of the toxicants onto them. The experiment was carried out over a four weeks period.
Trial 2 Due to the low recovery obtained from the PAH accumulation in roe (trial 1), and in light of the major oil spill that occurred along the Lebanese coast in 2006 whereby 15 000 tons of fuel oil covered a good surface of the sea urchin floor habitat), an additional bioaccumulation experiment was set up (trial 2) using some of the first trial toxicants (PAHs) but with a much higher concentrations and some additional metals (Nickel and Vanadium known to occur in fuel oil) as listed in table 1. Trial 2 toxicants were prepared using the same protocol as trial 1 but with two treatments: Control and Toxicants. The concentration of all toxicants used in this trial was 1000 ug l-1 (except for Anthracene 3000 ug l-1) with exposure lasting only two weeks. Concentrations for trial 2 were chosen from the percentage estimation of PAH in the crude oil based on the analysis of the Cedre laboratory report in France (Guyomarch 2006) . This percentage was multiplied by the 15 000 tons and by the volume of sea water (estimated to be 10 m in depth, 100m stretch seaward and over a hundred meter shore length) and are also closed to concentrations used in a previous study (Regoli et al. 2003). The tanks water was also tested one hour following of the addition of toxicants. For the metals analysis of this trial, the three replicates of the control treatment were pooled together in order to reduce costs (figure 6 of control treatment presented without ±SE).
Sampling and Analyses The first bioaccumulation experiment (trial 1) was set up over a 28-day incubation period. At days 0, 7, 14, 21 and 28, three sea urchins were randomly removed from each tank (referred to as week 0 to 4). The second bioaccumulation experiment (trial 2) was set up over a period of two weeks. At days 0, 7 and 14, three sea urchins were randomly removed from each tanks (referred to as week 0, week 1 and week 2).The sea urchins were rinsed with 1% nitric acid and acetone before being further processed. The diameter, total weight, roe weight and other body characteristics (color) were noted for each sampled sea urchin. Sea urchins were excised and
53 Chapter 3.Bioaccumulation separated into the following body parts: Roe, spines, teeth and test. In the first trial, no attempt was made to measure toxicants in the faeces, coelomic fluid, or the gut. In the second trial however, coelomic fluid, faeces and algae (sea urchin feed) were measured for some toxicants (Table 7). The roe was extracted using stainless steel or polyvinyl chloride (PVC) utensils. All samples collected from the same tank within each treatment group were pooled together. A sub- sample (0.2g) of the roe was taken from each treatment on the first week only (due to insufficient sample mass) for microbiological analysis before freeze-drying and then homogenizing the remaining roe samples. The roe samples were then used to determine the organic and inorganic loads. Spines, teeth and test were oven dried at 60 oC with each body part homogenized separately and tested only for metal bioaccumulation.
Metal analysis for trials 1 and 2 Metal analyses, for both trials, were done according to the protocol described in the temporal variation chapter. In summary, the body parts of sea urchins were digested in a microwave and run on ICP-MS.
Quality Control for metals The quality control protocol was used for both the digestion and the instrument analyses were also described in the temporal variation chapter.
The generated data was treated for both the digestion and analyses and calculated from both the external standard curve and the ratio of internal standard (IS) using Chemstation (B.03.07). Analytes‘ concentrations were corrected for extraction efficiency by multiplying the average recovery of the surrogate/internal standards by each analyte‘s concentration detected in the sample. Data in this experiment was further treated to adjust for lead concentration in roe (25% of data) when lead concentration of CRM was outside the certified limit and results were expressed as microgram of metal per g of dry weight (µg g-1). For statistical purposes values below the detection limit were reported as half the limit of quantification. The limits of quantification for metals are 1 µg l-1 in water and 0.01 µg g-1 in tissues. All methods used for digestion and analysis were sourced from internationally approved methods, however with some modifications (EPA-USA-200.8 1994, EPA-USA-3052 1996, APHA 1999).
Organic methods OCP-PAH
Microwave Extraction for trial 1
Sample extractions were performed using an Ethos 1 (1000 W) microwave extraction system (Milestone) equipped with a stirrer, temperature sensor and 10 vessels with a nominal volume of 100mL. Temperature and pressure inside the extraction vessels can reach a maximum of 260 °C and 35 bar (500 psi), respectively. The glassware was washed with detergent, rinsed with DI
54 Chapter 3.Bioaccumulation water and acetone (laboratory-grade, Fisher) before being heated in an oven at 180 oC for 2h. 0.1 g of homogenised roe was weighed and put in a glass extraction vessel designed for organic extraction. Surrogate standards Endosulfan II (99.5% certified, 1 mg l-1) and phenanthrene d-10 (1 mg l-1 certified) were added directly to the roe, mixed, and left to settle for half an hour. 3 ml of methanolic potassium hydroxide, and 10ml hexane-acetone (1:1) were added to the roe in the vessel before closing it for digestion. In preparation for column clean up, the final solvent extract was filtered and concentrated to 1ml using the Ethos post-extraction solvent evaporation apparatus equipped with a vacuum pump. All solvents used were of chromatography grade and purchased from suppliers such as Merck, Fisher and Supelco.
Clean up for trial 1
As a clean-up step, silica, florisil and sodium sulphate (2:1:2) were prewashed separately with dichloromethane (DCM) and left in an oven for 16 hours at 250˚C. The silica was then inserted in a glass column containing glass wool at the bottom, followed by florisil. Finally, the sodium sulphate was added with a drop of water to activate it. This clean up column was then conditioned with DCM, then Hexane-DCM (4:1) and finally hexane. The 1 ml extract from the extraction was added to the clean up column and left to bind without applying any vacuum. Before the last drop fell below the meniscus line, the extract was eluted (without vacuum) with hexane (for hydrocarbons), then Hexane-DCM (4:1) (for OCPs) and finally with DCM (for PAHs). The 3 fractions of the eluent were then pooled and combined into one fraction then evaporated till near dryness (100μl). 2ml of ethyl acetate was then added, and this evaporation process was repeated twice. The final volume of 0.5ml was filtered using 0.20 μm polytetrafluoroehtylene (PTFE) membrane filter before being transferred to a chromatography vial to be injected for analysis on Gas Chromatography Electron Capture Detector (GC-ECD) for OCP and on gas chromatography- mass spectrometry ( GC-MS) using scan and select ion monitoring (SIM) modes for PAH analysis and for PAH and OCP confirmation.
Extraction for trial 2 PAH
For this trial, two newly developed methods were used, with some modifications, to determine the PAHs in fatty matrices. The first method was adopted by the Association of Official Agricultural Chemists (AOAC) and is known as the QuEChERs method. The second method was developed by Ramalhosa‘s (2009). These two methods were chosen because they are less labor-intensive and use less solvent whilst yielding better recovery in difficult fatty matrices like roe. The extraction was preceded by a saponification step undertaken on 0.1 g of freeze-dried roe. For saponification, roe was placed in a glass tube (with Teflon stopper) and activated with a few drops of water. Roe was spiked with 1 mg l-1 of 99.5% chrysene d-10 as internal standard. 3 ml of methanolic potassium hydroxide and 1 ml of methanol were added to the spiked sample
55 Chapter 3.Bioaccumulation before placing it in a water bath for 1 hour at 80oC. After saponification, the sample was left to cool before adding 5 ml of deionized water and 10 ml of acetonitrile saturated with hexane to the saponified solution. The solution was then vortexed and the upper layer was transferred to 50 ml Teflon QuEChERs tubes using a Pasteur pipette. 4g magnesium sulfate anhydrous, 1g sodium chloride, 1g trisodium citrate dehydrate and 0.5g disodium hydrogen citrate sesquihydrate (Fluka) were added to the solution, vigorously shaken for 3 min, then centrifuged for 10 min at 3200 rpm. The supernatant was transferred to another 50 ml Teflon tube for the clean up step.
Clean up for trial 2 PAH
1200mg magnesium sulfate, 400mg C18, and 400mg primary secondary amines (PSA) (Bondesil) were added to the solution in the Teflon tube, vigorously shaken for 1 min, centrifuged for 10 min at 3200 rpm. The supernatant was evaporated to a 0.5ml, filtered using 0.20 μm PTFE membrane filter before being analyzed using high performance liquid chromatography (HPLC).
Analysis of OCP and PAHs for trial 1 All instruments used in this study had undergone an operational qualification and performance verification (OQPV). On the day of use, the chromatography instruments were tuned, maintained and checked for good performance using certified quality standards purchased from Agilent. All internal and external standards used for calculating response factor for PAHs and OCPs have a purity of more than 99% (confirmed by certificates of analysis when purchased from suppliers). Results of roe are not expressed in µg of organic toxicant g-1 of lipid but rather in µg of organic toxicant g-1 of dry weight of roe. This is due to the limited amount of roe available for the analysis at hand. The PAHs detection limits were 0.1 µg l-1 in fluids and 0.02 µg l-1 in water. For the microwave digestion, the PAHs limit of detection was 0.5 µg g-1in tissues and 0.1 µg g-1 for the QuEChERs method. . The OCPs detection limits were 0.01 µg g-1 in fluids and 0.02 µg l-1 in water. For the microwave digestion, the OCPs limit of detection was 0.005 µg g-1in tissues. Analytes‘ concentrations were corrected for extraction efficiency by multiplying the average recovery of the surrogate/internal standards by each analyte concentration detected in the sample.
OCP analysis Quantitative measurements were conducted using a Hewlett Packard (USA) 6890 Series GC with a 63Ni Electron Capture Detector (ECD). This was equipped with a JW-DB-608 column 32m length x 0.32mm internal diameter x 0.50µm film thickness. Split-Splitless injection of a 1 µl sample was performed by an HP G1513A auto-sampler with a 1 min solvent split time.
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The column temperature was programmed from 40°C to 160°C at 20°C/min, then further increased to 275 °C at 3°C/min then to 290 °C at a rate of 20 °C/min, flow rate was 1ml/min (the total run time was 55.08 min). The carrier gas was ultra high purity nitrogen at 18.4 ml/min. The injector and detector temperatures were both maintained at 280°C. Data was acquired and processed using Chemstation software (B.01.03 /SR1) compatible with an IBM PC and series interface unit.
OCP confirmation and PAH analyses and confirmation Peak identification and confirmation for both organic categories were performed on the samples using a HP 6890 Series GC-MS 5975c series equipped with an AT-5MS (30 m length, 0.25 mm diameter and 0.25 µm thickness) column, G1513A autosampler, operated at the electron impact mode (EIM) and using Chemstation software (D.03.00) for quantification. Mass spectra were acquired using both the scan and the SIM modes In the scan mode, due to the complicated chromatographic features associated with the roe complex matrix, components were identified by comparison of their mass spectra with the Wiley mass spectra library (version 1.4) and their retention time with pure standards. In the SIM mode, three fragmentation ions (when possible as some compounds have less than two) for each toxicant were chosen for mass scanning. The dwell time was 50 ms for windows. Temperature program was as follows: from 40 °C to 160 °C at a rate of 5°C /min then to 250 °C at a rate of 3°C /min and finally to 300 °C at a rate of 25 °C /min and constant for 10 min (total run time was 70min) Split-splitless mode, injection inlet at 250 °C Helium gas with a total flow of 18 ml/min, MSD transfer line at 300 °C, MS Quad 150°C and MS source 230 °C.
Analysis of PAHs for trial 2 Separation of analytes was performed using a C18 column ( PAH, 150 × 4.0 mm; 5 μm particle size) or equivalent assembled on a HPLC1100 system comprising of an autosampler, a binary pump, a degasser, temperature controlled column compartments, a diode array detector (DAD) and a fluorescent detector (1046A) . At room temperature (20oC ±2oC), an injection of 100 μL was made using an autosampler. The mobile phase consisted initially of 50% acetonitrile (ACN) and 50% deionized water. A linear gradient was used to increase to 100% ACN over 15 min. The flow rate was 0.8 mL/min. The initial conditions were achieved in 1 min and maintained for 5 min before the next run. The total run time was 40 min. The fluorescence wavelength program for each compound was detected at its optimum excitation/emission pair. The confirmation for this trial for the determination of PAH was done on GC-MS using the same program as that of trial 1. Chemstation software (B.03.02 SR1,SR2) was used for quantification.
Organic Quality control To evaluate both the extraction efficiency for the target compounds and the efficiency of the GC–MS for analysis of these compounds, recovery studies were carried out using both standard
57 Chapter 3.Bioaccumulation addition and an internal standard. The PAHs in the samples were identified by a combination of a retention time match and mass spectra match against the external calibration standards. Quantitation was performed by internal standardization using isotopically labelled phenanthrene-d10 and chrysene-d10. External standards of PAHs and OCP were used to calculate response factors by comparing their peak areas with the internal standard
Chemstation software was used for data acquisition. During the analysis, quality control (QC) was carried out on each set of samples to check performance of digestion (blank of digestion, spiked blanks, duplicates, sample spiked) and also to check performance of instrument (tuning, Ion ratios, standard curve, continuous calibration standard) with each batch of solvents, before and after maintenance of column and/or instrument consumables (washers, septum etc.).
Microbiological methods Within this study, fecal coliforms were used as indicators of microbiological accumulation. The method used for the determination of fecal coliform was adopted from the ISO and APHA . All tools used in microbiology analyses were autoclaved at 123oC for 30 min. The gonads samples from each treatment (0.1g) were homogenized using a stomacher bag after adding10 ml of 10% peptone water, a medium known to enhance growth of total coliforms and E.coli. 1ml of that suspension was transferred to an m-endo agar for the growth of total coliform and another 1ml was transferred on m-FC agar for the growth of fecal coliform. Incubation took place at 35 oC for 24 h for m-endo agar and 24 h at 44 oC for the m-FC agar before identification and confirmation were performed biochemically. The procedure for the confirmation of coliforms is as follow: Lauryl tryptose broth and brilliant green bile broth were inoculated by transferring growth from each colony of the m-endo agar. Gas production in both tubes that have been incubated at 35C±0.5 oC for 24 h verified the idendity of coliforms. Fecal coliforms were confirmed from a total coliform agar sample by swabbing the colonies with a cotton swab and inoculating a tube of EC medium broth at 44C ±0.2oC for 24 h. Gas production in the EC medium confirmed the presence of fecal coliforms. All chemicals used for confirmation were purchased from HACH.
The Bioconcentration Factor The bioconcentration factor, expressed as the quotient of the concentration of a chemical in aquatic organisms at a specific time or during a discrete time period of exposure divided by the concentration in the surrounding water at the same time or during the same period (ATSDR 2008), was calculated for both trials by subtracting the control value from the calculated concentration and dividing the calculated concentration from the bioaccumulation experiments with the nominal value of the given concentration for each week. It was therefore expressed
58 Chapter 3.Bioaccumulation with no units (BCF= (Sample concentration in tissues- control concentration in tissues) (in µg g- 1 of dry tissue) / Nominal concentration: in rearing medium .mg l-1). It is assumed that the majority of bioaccumulation happened via the water column.
Data treatment and statistical analyses For trial 1, a Repeated Measure ANOVA (significance at 0.05) was used to detect differences in bioaccumulation of sea urchins with weeks as ―within factor‖ and body parts and treatments as ―between factors‖ (fixed). For the organic trial 2, a repeated measure ANOVA was used with weeks as ―within factor‖ and treatments as ―between factors‖ testing for difference in the roe of the organism. For both trials Normality assumption was tested via a Q-Q plot (Quinn et al. 2002). Sphericity was tested using Mauchlan‘s test. When sphericity was violated, both Greenhouse-Geisser epsilon p values and Greenhouse-Geisser modified degree of freedom were considered instead. Data was square root transformed for the repeated ANOVA in both trials. For Treatments and body parts post hoc tests, a Tukey (α=0.05) was used and appended to the ANOVA tables with means of concentrations ranking by decreasing order from left to right. Terms underlined indicate that they were not significantly different. The post hoc for weeks was interpreted from the graph. SPSS (2010) software (version 18), compatible with IBM, was used for all statistical analyses.
For both trials, I sought to check if size was to be considered a cofactor to include in the bioaccumulation statistical analysis. Accordingly, a One Factor ANOVA was used for each of the four weeks separately to check for differences in the size, total and roe weight of the sea urchins with treatments (control, half and full) as factor. I did not conduct any Repeated ANOVA to test for the differences in size as I expected the growth of sea urchins between beginning and end of experiment and also because I was not interested in the interaction between days and treatments, although difference in sizes between treatments within each week could potentially cause problems. For trial 2, one factor ANOVA was also used for each of the two weeks with treatments (control and toxicants) as factors. Normality test (Q-Q plot) and Homogeneity (Levene test) were conducted before the use of ANOVA.
Results
First trial
Size, Total Weight and Roe Weight Within each of the four weeks of the first trial, the sea urchins were of similar size and roe weight, which eliminates these parameters as cofactors affecting bioaccumulation between
59 Chapter 3.Bioaccumulation contaminant treatments (Figure 1). The difference in total weight (less than 25%) between treatments in week 4 is possibly due to a loss of coelomic fluid when the sea urchins were temporarily stored in fridges overnight before being processed the following day.
Figure 1. Average total weight (g ±SE , size (test diameter in cm ±SE), and roe weight (g ±SE) of the sea urchin Paracentrotus lividus used for the control, half and full treatments over four weeks of the bioaccumulation experiment in the first trial. Results from the ANOVA significance p (α=0.05) are encased. Only total weight in the fourth week was significantly different perhaps due to the loss of some coelomic fluid.
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Lead and Cadmium in tissues The accumulation of cadmium and lead in the different body parts of the sea urchin increased with time in the contaminant treatments with the roe scoring the highest accumulation among all body parts and the teeth scoring the lowest. For both metals, there was a strong interaction between treatments, body parts and weeks. In general, the control treatment showed no bioaccumulation of cadmium and lead in all body parts except for an increase in cadmium in spines observed in the first two weeks.
The bioaccumulation of cadmium in the roe increased over the four weeks for both the half and full treatments with the full treatment (6.6 µg g-1) reaching double the concentration of the half treatment at the end of the fourth week. For test and spines, the bioaccumulation was slow in the first 2 to 3 weeks. However, in the final week, both half and full treatments reached similar concentrations in the test. Whereas in spines, metal bioaccumulation in the full treatment was double that of the half treatment. The lowest bioaccumulation occurred in the teeth, where the full treatment reached 0.3 µg g-1 at the end of the 4th week (Figure 2, Table 2).
Lead was found to accumulate in the roe, spines and teeth. Its accumulation increased with weeks of exposure, where the concentration in the full treatment reached double that of the half treatment. Teeth accumulated the least lead. Only in test, half and full treatment scored close values at the end of the fourth week, showing similar results to the bioaccumulation of cadmium (Figure 3, Table 2) by the sea urchin.
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Figure 2. Bioaccumulation of cadmium (µg of metal g-1 of dry weight tissue ±SE) in different body parts (Roe, Test, Spines, Teeth) of sea urchin Paracentrotus lividus over four weeks of exposure to three treatments (control, half and full). Week 0 is the baseline before the start of experiment.
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Figure 3. Bioaccumulation of lead (µg of metal g-1 of dry weight tissue ±SE) in different body parts (Roe, Test, Spines, Teeth) of sea urchin Paracentrotus lividus over four weeks of exposure to three treatments (control, half and full). Week 0 is the baseline before the start of experiment.
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Table 2. Repeated Measure ANOVA testing for differences in bioaccumulation of cadmium and lead in different body parts of sea urchin Paracentrotus lividus with weeks as factor within subjects and treatments and body parts as factors between subjects. The multiple comparison tests of Tukey performed after Repeated Measure ANOVA are summarized in the appended table. Mean metal concentrations are ranked from the left to the right by decreasing order. Terms in underlined compartments, body parts or treatments are not significantly different (αTukey= 0.05).All data was square root transformed. When Sphericity assumption was violated the Green house Geisser epsilon adjusted p values and an adjusted degree of freedom were used.
Cadmium a df MS Sig Between Subjects Treatment 2 2.707 0.000 Body parts 3 5.505 0.000 Treatment*Body parts 6 0.579 0.003 Error 24 0.124 ---
Within Subjects Weeks 4 2.306 0.000 Weeks*Treatment 8 0.416 0.000 Weeks*Body parts 12 0.351 0.000 Weeks*Treatment*Bodyparts 24 0.103 0.001 Errors (Weeks) 96 0.040 Lead a,b df MS Sig Between Subjects Treatment 2 2.327 0.000 Body parts 3 2.193 0.000 Treatment*Body parts 6 0.211 0.001 Error 24 0.036 ---
Within Subjects Weeks 2.29b 1.594 0.000 Weeks*Treatment 4.579b 0.484 0.000 Weeks*Body parts 6.869b 0.800 0.000 Weeks*Treatment*Bodyparts 13.737b 0.071 0.025 Errors (Weeks) 54.949b 0.033 --- a All data has been square root transformed b Denotes use of Greenhouse-Geisser epsilon adjusted p-values and degree of freedom adjusted values due to assumption of sphericity violation (when G-G epsilon is < 0.7)
Cadmium Post Hoc (week 4) Lead Post Hoc (week 4)
Full Half Control Full Half Control
Roe Spines Test Teeth Roe Test Spines Teeth
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Dieldrin and DDE in roe The level of bioaccumulation of dieldrin and 4,4‘DDE in the roe of P. lividus showed a different pattern to the metals. While the bioaccumulation of metals increased until the end of the fourth week, the bioaccumulation of dieldrin and 4,4‘DDE appeared to reach a plateau by the second week (Figure 4, Table 3).
For dieldrin bioaccumulation, there was no significant interaction between weeks and treatments. The bioaccumulation in the roe at the end of the experiment (4 weeks) of half and full treatment ranged from 100 to 191 ug g-1.
For 4,4‘DDE bioaccumulation, there was significant interaction between weeks and treatments. For the first three weeks, the bioaccumulation of the full treatment was at least 1.5 fold greater than the half treatment. However, on the fourth week, the half treatment attained a slightly higher concentration (457 µg g-1) than the full treatment.
Fat content of 12 sea urchins randomly collected from both trials was 13.2% (SE±0.05).
Figure 4. Bioaccumulation of dieldrin and 4,4‘DDE (µg of pesticide g-1 of dry weight tissue ±SE) in the roe of sea urchin Paracentrotus lividus over four weeks of exposure to three treatments (control, half and full). Week 0 is the baseline before the start of experiment.
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Table 3. Repeated Measure ANOVA testing for differences in bioaccumulation of dieldrin and 4,4‘DDE in the roe of sea urchin Paracentrotus lividus with weeks as factor within subjects and treatments as factor between subjects. The multiple comparison tests of Tukey performed after Repeated Measure ANOVA are summarized in the appended table. Mean pesticide concentrations are ranked from the left to the right by decreasing order. Terms in underlined compartments, body parts or treatments are not significantly different (αTukey= 0.05).All data was square root transformed. When Sphericity assumption was violated the Green house Geisser epsilon adjusted p values and an adjusted degree of freedom were used.
Dieldrin a,b df MS Sig Between Subjects Treatment 2 100.059 0.009 Error 24 5.243 ---
Within Subjects Weeks 1.216b 61.833 0.137 Weeks*Treatment 2.432b 30.684 0.306 Errors (Weeks) 4.864b 19.649 --- 4,4‘DDE a df MS Sig Between Subjects Treatment 2 438.803 0.011 Error 4 25.183 ---
Within Subjects Weeks 4 128.43 0.000 Weeks*Treatment 8 68.721 0.003 Errors (Weeks) 16 13.705 a All data has been square root transformed b Denotes use of Greenhouse-Geisser epsilon adjusted p-values and degree of freedom adjusted values due to assumption of sphericity violation (when G-G epsilon is < 0.7 4,4‘DDE Post Hoc Dieldrin Post Hoc
Full Half Control Full Half Control
Acenatphthene and Pyrene in roe Acenaphthene and pyene accumulated in the roe of the sea urchin P. lividus. However this bioaccumulation could not be properly quantified due to some limitations in the detection limit of the method used in trial 1 and to the complexity of the matrix. It was not possible to report it between instrument detection limit and quantification limit because the outcomes of the results were not replicated enough within each treatment. Another trial was set for these compounds increasing the spiking level in each treatment by 100 fold (trial 2) in light of the oil spill.
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Second trial
Size, Total Weight and Roe Weight Within each of the two weeks of the second trial, the sea urchins were of similar size and roe weight, which eliminates these parameters as cofactors affecting bioaccumulation (Figure 5).
Figure 5. Average size (diameter in cm ±SE), total weight (g ±SE) and roe weight (g ±SE) of the sea urchin Paracentrotus lividus used in the control and treatment over two weeks of exposure in the second trial. The sea urchins were kept one additional week for the sole monitoring of size, total weight and roe weight till the death of all sea urchins. Also reported within the graphs are the p values (significance 0.05) of One Factor ANOVA with treatments (control and toxicant) as factor.
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With regards to the bioconcentration factor, it is important to note that the bioaccumulation of cadmium and lead in the different body parts of the sea urchin P.lividus differed from that of trial 1. In general, by the end of the experiment, roe had the highest accumulation for cadmium and lead (Figure 6).
The cadmium bioaccumulation occurred in all body parts, but was the highest in the roe. After an initial depuration, roe had accumulated up from 16 to 26 µg g-1 at the end of the second week. Also at the end of second week, test had accumulated 1.5 fold the concentration of first week (9 and 13 µg g-1). Whereas in spine, the bioaccumulation at the end of second week was double the accumulation of the first week (4.2 and 8.8 µg g-1). There was a tenfold increase of cadmium bioaccumulation in the teeth on the first week, which remained constant by the end of the second week (Figure 6).
Over the two weeks exposure period, lead has accumulated in the roe and spines more than in test and teeth. It increased from 0.11 to 1.06 µg g-1 in roe and from 0.18 to 1.28 µg g-1 in spines. In test and teeth, it increased from 0.06 (test) and 0.13 µg g-1 (teeth) in control to 0.15 (test) and 0.32 µg g-1 (teeth) in spiked treatment during the first week (Figure 6).
Vanadium was not taken up by the roe, test or spines. An increase by 1.5 fold was seen in the teeth (from 0.24 µg g-1 to 0.38 µg g-1) by the end of the second week in spiked treatment. The nickel in the roe slightly increased in the first week and dropped back to almost original levels on the second week. Test and spines increased slightly but not significantly by the end of the second week as for the teeth of the sea urchin a slight increase was detected with no significance at the statistical level.(Figure 6).
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Lead, Cadmium, Nickel and Vanadium in tissues
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Figure 6. Trial 2 bioaccumulation of lead, cadmium, nickel and vanadium (µg of metal g-1 of dry weight tissue ±SE) in different body parts (roe, test, spines, teeth) of Paracentrotus lividus over two weeks of exposure to two treatments (control and toxicants). The control samples are presented without SE as they were composited from three replicate tanks.
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Anthracene, pyrene and phenanthrene in tissues In this trial, the roe concentration by the end of the second week almost doubled the concentration of the first week with regards to phenanthrene and pyrene. For anthracene, I did not get good recovery for the first week, however, in the second week, the roe accumulated 58 µg g-1of anthracene PAHs were also recovered in the faeces and the coelomic fluid (Figure 7 and Table 7).
Figure 7. a-Bioaccumulation of phenanthrene, anthracene and pyrene (µg of organic g-1 of dry weight tissue ±SE) in the roe of sea urchin Paracentrotus lividus over two weeks of exposure to two treatments (control, spiked toxicant), b-accumulation of the three PAHs in the roe, faeces and coelomic fluid at day 14 of trial 2.
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Bioconcentration factor and metal concentration ratio in sea urchins body parts In trial 1, the metal BCF in roe from the half treatment was higher than the accumulation from the full treatment. At the end of the 4th week, the metal concentration ratio of full to half was equivalent to 1, which shows a linearity of uptake (Table 4). Roe was the highest accumulator of metals. As for OCPs, the same trend as in metals was seen where the urchins from the half treatment bioaccumulated more than the urchins from the full treatment. There was no prediction of linearity, as a plateau was reached after the second week.
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Table 4. Bioconcentration factor in sea urchins‘ roe, test, spines and teeth (dry weight) for cadmium, lead, 4,4‘DDE, and dieldrin for trial 1 based on dry weight. A ratio of full treatment BCF to half treatment BCF gives an indication of linearity in uptake ( the closer the ratio is to 1, the linear the uptake is). Third column is the ratio of body parts to each other for both treatments at the end of the experiment. BCF was calculated according to this formula: BCF = (concentration in tissues – Concentration in
Control) µg g-1/ Nominal concentration of medium µg l-1 .
First trial Bioconcentration factor (BCF) Ratio of Full/Half treatment Metal Ratio in Body parts at week 4 Roe Roe Test Test Spines Spines Teeth Teeth Roe Test Spines Teeth Roe:Test:Spines:Test
Cadmium full half full half full half full half Full/Half Full/Half Full/Half Full/Half Half treatment Cd Week 1 83 (±30) 197 (±136) ------16 (±4) 12 (±4) 23 (±21) --- 0.42 --- 1.33 --- 18:8:4:1 Week 2 366 (±19) 531 (±190) ------49 (±13) 28 (±3) 9 (±4) 24 (±4) 0.68 --- 1.75 0.375
Full treatment Cd Week 3 156 (±22) 270 (±42) ------69 (±11) 77 (±9) 6 (±2) 19 (±15) 0.57 --- 0.89 0.31 22:5:6:1 Week 4 352 (±100) 400 (±99) 10 (±4) 30 (±31) 151 (±15) 140 (±15) 49 (±13) 49 (±12) 0.88 0.33 1.07 1
Lead Half treatment Pb Week 1 146 (±120) 75 (±33) 26 (±2) 13 (±14) 52 (±15) 36 (±18) 16 (±7) 29 (±19) 1.94 2 1.44 0.55
4.7:1.9:1.76:1 Week 2 340 (±32) 406 (±104) 52 (±4) 29 (±24) 71 (±13) 92 (±30) 13 (±8) 26 (±5) 0.83 1.79 0.77 0.5
Full treatment Pb Week 3 515 (±165) 391 (±61) 23 (±3) 26 (±10) 164 (±19) 173 (±23) 14 (±1) 33 (±7) 1.31 0.88 0.94 0.42 5.2:1.3:2.3:1 Week 4 637 (±148) 703 (±121) 117 (±13) 251 (±75) 178 (±19) 147 (±2) 29 (±4) 39 (±4) 0.90 0.466 1.21 0.74
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Table 4. continued
BCF-Full BCF-Half Full/Half
4,4'DDE Week 1 13674 (±4893) 7475 (±1360) 1.8
Week 2 15267 (±NA) 22971 (±7645) 0.7
Week 3 16676 (±2586) 17933 (±9710) 0.9
Week 4 14342 (±7916) 36604 (±11702) 0.4
Dieldrin
Week 1 5973 (±1410) 3320 (±342) 1.8
Week 2 4938 (±18) 7882 (±1736) 0.6
Week 3 5904 (±942) 6572 (±3733) 0.9
Week 4 7348 (±3987) 7096 (±3784) 1.03
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In trial 2, the BCFs for cadmium and lead were lower than trial 1 despite the fact that the exposure concentration was 100 fold higher. The few points for the PAHs do not allow a judgement of the BCF (Table 5).
Table 5. Trial 2, BCF in roe of sea urchins (dry weight) for cadmium, lead, nickel, vanadium, phenanthrene, anthracene and pyrene in weeks 1 and 2 after exposure to 2 treatments (control and toxicants). BCF = (concentration in tissues – Concentration in Control)µg g-1/ Nominal concentration of medium µg l-1 .
Trial 2 BCF in Roe BCF inTest BCF in Spines BCF in Teeth Cadmium Week1 13.2 (±2) --- 4 (±0.1) 4 (±0.7) Week2 26 (±2.5) 12 (±6) 8 (±2.3) 4 (±0.1) Lead Week1 0.5 (±0.08) 0.1 (±0.02) 0.3 (±0.05) 0.2 (±0.02) Week2 1 (±0.1) 0.3 (±0.07) 1 (±0.1) 0.4 (±0.02) Nickel Week1 1.86 (±0.9) --- 0.3 (±0.1) 0.3 (±0.2) Week2 --- 3.07 (±1.5) 0.4 (±0.5) 0.2 (±0.02) Vanadium Week1 ------0.14 (±0.05) Week2 ------0.14 (±0.03) Ratio of week2 BCF to week 1 BCF Phenanthrene Week 1 357 (±25) ------Week2 894 (±91) 2.5 ------Ant hracene Week1 ------Week2 19 (±2) ------Pyrene Week1 205(±19) ------Week2 362 (±60) 1.76 ------
In table 6 below, tabulated measurements of pH, Dissolved oxygen, conductivity and temperature taken for the water tanks during experiments and averaged (Count: 135) for both trials are presented.
Table 6. Mean (±SE) of physical and chemical parameters monitored in the tanks during trial 1 and 2: pH, -1 -1 o dissolved oxygen (mg l ), conductivity (mS cm )/salinity %o (part per thousand) and temperature in C.
Parameters pH Dissolved Conductivity/Salinity mS/cm Temperature in oC -1 oxygen* mg l / %o Trial 1 8.2 (±0.02) 8.2 (±0.02) 42.3 (±0.045) 22.9 (±0.06) Mean(±SE) Trial 2 8.1 (±0.01) 6.21 (±0.01) 42.2 (±0.1) 22.4 (±0.02) Mean(±SE)
* New Dissolved Oxygen meters was used for trial 2
75 Chapter 3.Bioaccumulation Table 7. Cadmium, lead, pyrene, anthracene and phenanthrene concentrations in water, algae (fed to sea urchin), faeces and coelomic fluid from trial 2 at different stages
of the experiment. Metals spiking levels for trial one was determined solely for cadmium and lead at the 5 and 10 µg/L.
Parameters First trial Cd Pb Pyrene Anthracene Phenanthrene Cadmium: 5 µg/L: 4.8µg/L, 10 µg/L: 8.9 µg/L, Lead: 5 µg/L: 2.8µg/L, 10 µg/L: 5.7 µg/L : Parameters Second trial Baseline value of sea water in tanks before spiking µg/L a 27.48 4.1 Not detected Not detected Not detected Nominal spiking concentration µg/L 1000 1000 1000 2000 1000 Water Actual concentration 1 hour after spiking µg/L 1386 637 492 756 Not detected Actual concentration after one week of spiking and before change of water µg/L 1053 122 Not tested Not tested Not tested -1 0.223 4.2 Algae b Sample from control treatment collected after 48 hs µg g Not detected Not detected Not detected Sample from spiked tanks collected after 48 h µg g-1 31.9 272 8.0 24.3 4.1 -1 Faeces Sample from control tanks collected after two weeks µg g 0.216 1.05 Not detected Not detected Not detected Sample from treatment tanks collected after two weeks µg g-1 203.7 93.8 314 45.2 229 -1 Sample from control collected at the end of experiment mg l 1 0.006 Not detected Not detected Not detected -1 Sample from treatment collected after one week of bioaccumulation mg l 1.2 0.133 Not tested Not tested Not tested Ceolomic Sample from treatment collected after two weeks of bioaccumulation mg l-1 6.0 1.254 83 12 103 -1 Sample from treatment collected after three weeks of bioaccumulation mg l 48.3 28.05 Not tested Not tested Not tested a-Suspected contamination in the bottle used to collect this sample after unwashed residues were found at the bottom of bottle. Anyway, the level is smaller relatively to the 1 mg l-1 of the spiked solution. b-Algae were kept in separate tanks, spiked with the same amount as sea urchins tanks and collected after 48 h to determine how much plant accumulated over two days which is the total time the sea urchins were fed during the bioaccumulation experiment.
Table 8. Bacteriological load, average of fecal coliform colony forming unit (cfu) per gram of sea urchin P lividus‘ roe trial 1 at the end of the first week for the three treatments.
Fecal coliforms cfu/g of roe Treatment Average Control > 6 x 106 colonies per 1g Half > 6 x 106 colonies per 1g Full > 6 x 106 colonies per 1g 76 Chapter 3.Bioaccumulation
The baseline of the bacteria load in the roe of the sea urchin P. lividus’ roe was not determined upon arrival to the laboratory. It is not clear if the level of colony forming units of fecal coliforms seen in the control (table 8) was loaded during the acclimatisation period (sea water not tested for coliforms) or back in the field. The addition of more bacterial culture did not cause any increase in bioaccumulation of these bacteria. This can be due to the fact that not enough serial dilutions were done, because toxicants may have killed the bacterial culture when the spiking mixture was prepared, or because the amount given is insignificant relatively to the current load. However, it is obvious that sea urchins have the capacity to take up microbes from their surrounding medium, a fact also confirmed by Portocali et al. (1996) and it is also known that the sea urchin can neutralize bacteria due to coelomocytes immune mechanism (Chia et al. 1996), but this may be linked to the concentrations of bacteria.
Discussion This study shows that the sea urchin Paracentrotus lividus effectively accumulates metals, PAHs and OCPs in its roe and accumulates metals in the test, spines and teeth. Moreover, for metals, this accumulation process varies between body parts in descending order (Roe> Test> Spines> Teeth). The variation was especially evident between roe and the other body components and between treatments (first trial). It is probable that the roe showed the highest uptake rate because it is the most metabolically active body organ out of the four body parts tested in the urchin. Moreover, this study took part during a period in which the roe was at its peak in terms of weight and activity (Byrne 1990). This is in agreement with Warnau‘s (1997) study, where the digestive wall and gonads showed the highest uptake for cadmium via a non dietary pathway (via water).
Metals
If the metal exposure is restricted to water borne, then the BCF may be used to evaluate the progress of uptake in this study. I have assumed that metals were largely taken up via seawater considering the urchins were only fed on one night at the end of 7 days (less than 24 hours) before the change of water occurred. If the uptake is linear, then it may be possible to extrapolate to other natural concentrations of P. lividus in the field. In this situation, I expect the ratio of BCF between the half and the full treatment to be equal to one. It was evident at the end of the study, that with the exception of the test, this ratio is indeed close to one. Warnau et al. (1997) reported the average BCF of cadmium in P lividus (based on exposure to 0.11 and 0.57 µg l-1 treatments over 24 days) to be 400 ±50 in the roe and 30 ±3 in teeth. Average cadmium BCF s in P. lividus in this study (based on exposure to 5 and 10 ug l-1 over 28 days) was 375 ±100 in the roe and 49 ±12 in the teeth. These levels are in agreement with Warnau. This is
77 Chapter 3.Bioaccumulation remarkable taking into consideration the gap of years between the two studies (change of techniques), the 10 fold higher concentration used in this study and the natural variation in the species across geographical locations.
In the first two weeks of exposure, I observed that for both metals cadmium and lead, the lower concentration resulted in a higher rate of bioaccumulation. This fact was also noted by Clason et al. (2003) who observed differences in bioaccumulation between high doses and low doses, suggesting that high contaminant concentrations can affect the biokinetics of metal uptake by the organisms, perhaps through saturation of binding sites. This is due to the fact that it has been proven that metal uptake across the biological membrane is protein-mediated (Veltman et al. 2010) among other pathways.
Rainbow (2007) stated that excess of metal at the cellular level must be detoxified, i.e. bound tightly to a ‗sacrificial‘ site from which escape is limited, most likely in a storage organ beyond the site of uptake. Saturation of binding sites (i.e. when no more binding with metals or sequestering with metallotioneins) in the roe (site of uptake) of P. lividus might explain why calcites such as the test, spines and teeth (considered here to be the sacrificial sites) did not start accumulating (with a noticeable increase in metal content) until the third week, while the teeth resisted taking up metals until the fourth week. This delay in the uptake of calcites to metals may also be considered a good defense mechanism by the sea urchin as its calcites‘ hard exterior serves as a means for the organism to resist attacks from its predators. According to Magdans (2004):‖ the size difference between Ca2+ and Mg2+ ions impedes the propagation of cracks, thus affecting the fracture behavior of the biogenic calcite. The accumulation of Mg in the spine-base most likely serves to enhance this effect and thus strengthens the spine close to the shell‖. Accordingly, when lead and cadmium replace the calcium-magnesium bond, the test, spines and teeth will soften rendering the sea urchin an easy target to its preys. In this case, redistributing the load of metals to calcites (relocation) might be a survival tool worth utilizing as a last resort, specifically in the case of saving the metabolic activity of the roe and avoiding permanent damage to the tissue. In summary, this implies that physiologically, metals are not accumulated in calcites first (a strategy perhaps that helps keep the calcites hard) and instead metals are bound in the roe up to a level where metals can no longer be contained /sequestered and become toxic to the roe at which point the physiological response would be to send metals to calcites to save the metabolic activity of the roe.
The role that the metallothionein plays in the defense mechanism of sea urchins is not very clear as the literature data present many contradictions and inconsistencies in the induction of these compounds (Amiard et al. 2006, Mao et al. 2012). Scudiero (Scudiero et al. 1994, 1995) has indeed isolated metallothionein from eggs and adults of P. lividus and also isolated other metal
78 Chapter 3.Bioaccumulation binding proteins (Scudiero et al. 1994) but still, little is known about the mechanism. How much metal will produce how much metallothionein is not quantified in sea urchins. These biomarkers can tell us if an exposure occurred but does not quantify this exposure. Moreover, Temara et al., (1997) reported that in another echinoderm, A. rubens, metallothionein was detected in pyloric coeca of the sea star but not in its gonad. In terms of defense, metallothioneins are not the only probable way of defense, the role of coelomocytes in the coelomic fluid is not to be ignored as these cells sequester toxicants before any damage to the DNA occurs (Loram et al. 2012, Matranga et al. 2012).
The size of the urchins within each week did not differ between treatments, which eliminated size as a cofactor in the bioaccumulation. In trial one, no change in roe weight was noticed at the end of the experiment. However with trial 2 (exposure to 1000 ug l-1), a drop in roe weight was noticed by one fold (but was ruled out to be significant by ANOVA) when the experiment was sustained till the third week (figure 5). In trial 2, roe accumulated lead and cadmium more than the other body parts, however, it did not accumulate the expected BCF as calculated in trial one. This was probably due to two reasons: either the binding sites were saturated, (as discussed above) or the organism was dying and signs of stress were showing. A large number of the sea urchins lost their spines on the third day. Some antagonist effects might have occurred here, as the organics and the metal compounds competed for carriers across the cellular membrane while the actual cell defends itself by detoxification. This is achieved through production of antioxidants and through carriers on the bi-lipid membrane, which have more affinity for organic compounds (size preferences) than for metals (Newman et al. 2008). It is important to note that the spines have accumulated the highest concentration of lead (more than roe, test and teeth), which can support the relocation theory, especially if the sea urchin might have used this method as a fast detoxification pathway, considering it would soon shed the spines and effectively excrete this contaminant load. This hypothesis is worth exploring since urchins in fact can regrow their spines once they are shed after trauma (Ebert 1967, Heatfield 1971).
The absence of vanadium and nickel uptake (trial 2) in the roe may indicate the metals were not bioavailable. This might occur if pH or oxygen level affected the speciation (see Table 6 for pH and oxygen levels in the tanks of both trials). However, it might also be that the organisms were struggling and under stress. Miramand et al. (1982) found that uptake of radiotracer vanadium was slow in P. lividus and that after 3 weeks of exposure isotopic equilibrium had not been reached. Similarly, Miramand and Fowler (Miramand et al. 1982) reported that Vanadium has shown weak penetration in the internal organs of benthic larvae. This fact was also mentioned by Venkataraman and Sudha (2005), who observed a poor absorption of Vanadium from the gastrointestinal tract in humans. Vanadium was originally recognised by its ability to inhibit
79 Chapter 3.Bioaccumulation membrane sodium pump and its capacity to affect the activities of various other intracellular enzyme systems (Venkataraman et al. 2005). In another kinetic experiment conducted by the authors of this study (Chapter 4), values of vanadium and nickel, tenfold lower, were detected in all body parts of the sea urchin but in low concentrations. In trial 2, the observed nickel and vanadium deposition in calcites was as seen in field in the temporal variation chapter, where it is believed that nickel and vanadium have more affinity to calcites than to soft tissue, regardless of the exposure route. Deposition of these metals in the calcites is usually affected by the length of exposure as seen in lead and cadmium.
Organics In trial 1, the organic compounds OCPs and PAHs, accumulated in the roe, and a plateau was reached after the first week for dieldrin and DDE. The latter compounds are lipid solubles, as reflected from literature by their log octanol-water partition coefficients Kow (Table 1). The BCF of both dieldrin and 4,4‘DDT (Table 4) exceeded the classification of the Stockholm convention on POPs and other chemical management programs which categorized chemicals with BCF higher than 5000 (wet weight basis) as bioaccumulative (Borgå et al. 2005). It is not known at this stage if uptake levelled off because of the saturation of the binding sites, a possible breakdown mechanism in the metabolism(denaturing proteins, reactive oxygen species (Novo et al. 2008 )) or detoxification (production of antioxidants) leading to a steady state.
4,4‘ DDT was spiked in the water but not recovered in the roe. It is believed that it was metabolized in the roe to 4,4‘DDE, a fact which has been observed in several DDT exposure studies (ATSDR 2002b). Gray (2002) reported that DDT and DDE accumulated in hawks and eagles resulting in the thinning of egg shells leading to the eventual death of the chicks. Ultimately populations were affected, as this substance acts indirectly on the calcium channel. When DDT is accumulated by sea urchins the scenario of calcites weakening can be expected. Similarly, if DDT and its metabolites are passed on to larvae, a whole generation of sea urchins can be affected by this uptake. Also, a combination of DDT, DDE, lead and cadmium might exhibit an additive effect weakening the calcite, a probability which is worth testing in future studies on the test, spines and teeth of the sea urchin along with the individual effects of DDE and cadmium. BCF values for DDE in the literature for benthic and fish ranged from 2,750 in sea urchins (Thompsons et al. 1989) to 4,550 in mussels to 690,000 in fish (ATSDR 2002b). As for BCF values for dieldrin in the literature, these ranged from 247 in crabs to 114,935 in snails (ATSDR 2002a). The average BCF ratio of full treatment (25 µg g-1) on the first week to half treatment (12.5 µg g-1) when uptake was almost linear is equivalent to 2, indicating there is a difference in toxicity at different levels of the treatments in the uptake.
80 Chapter 3.Bioaccumulation
Acenaphthene and pyrene were not detected with good replication by the end of the first trial perhaps due to some limitations in the extraction method and the complexity of the freeze-dried roe matrix. Acenaphthene in particular, might not have been taken up efficiently due to its low Kow. Another explanation might be that the sea urchins, having similar detoxification mechanism to humans as reported by Sugni et al. (2010), were eliminating the PAH when taken up in small doses (PAHs are biotransformed by the P-450 enzymes known to be present in P. lividus) (Den Besten 1998, Snyder 2000). This event was also noted in sea star Asteria rubens another echinoderm, which produced a benzo-a-pyrene hydroxylase formed in microsomal fractions of the pyloric caeca 3 to 4 days after injection of benzo-a-pyrene a compound of the family of PolyAromatic Hydrocarbons (Den Besten et al. 1993, Den Besten et al. 2001). It is worth noting that metals and PAHs were taken up by the algae and might have biomagnified in sea urchins even if the sea urchins were only exposed to algae for less than 14 hours. This period of 14 hours seemed to be enough for the algae to accumulate toxicants.
The range of BCF obtained in this study for the second trial conducted following the 2006 oil spill off the Lebanese coast, (1mg l-1 for pyrene and phenanthrene and 3 mg l-1 for anthracene ) falls within the middle of the range reported by literature for benthic and fish 10- 10,000 (ATSDR 1995, Baussant et al. 2001). In this second trial, behavioural and physiological changes were observed in the sea urchins P. lividus upon the addition of toxicants to the tanks. In all three tanks, sea urchins P. lividus were no longer attached via their tube feet to any substrate like walls of the aquaria or the filter. They detached and fell 10 seconds after the addition of the spikes but did not die. Tube feet have a high adhesive strength (force per unit area) due to some secretory cells with dual functions: secreting adhesive material to attach and de-adhesive material to detach (Santos et al. 2009, Santos et al. 2013). It is possible that the toxicants acted in one of three ways: a) they stimulated the secretory cells to produce de- adhesive, b) they directly killed the secretory cells (this would have been too fast) or c) the mixture of toxicants simply acted on the adhesive itself (a most probable scenario) diluting the adhesive hence weakening the contact point. Covering behaviour (by any shells or algae), still seen in control tanks, was not expressed in the toxicant tanks. Spines of urchins in toxicant tanks, were shed on the third day and the clumping of sea urchins in groups was no longer observed, as they appeared to be dispersed (Figure 8). The sea urchin showed great signs of stress, succumbing to fungal and bacterial infections, eventually leading to the death of some urchins starting the third week. This has slightly affected the roe weight as mentioned above. Perhaps this was due to loss of lipid (by not being able to eat and consume their reserve) as means of purging some of the organic load or simply because PAHs at high concentration might overcome the ability of the cell to detoxify by metabolizing. This suspected collapse in the metabolic pathways may also have affected the metal uptake in the second trial. One might
81 Chapter 3.Bioaccumulation suggest that an additive effect (to the stress) may have occurred. This is a probable scenario, which appears to have mimicked the effects on sea urchins inhabiting the Mediterranean Sea floor off the Lebanese coast, when the crude oil was released into their habitat in 2006.
Determining the tissue body burden remains the ultimate goal of many exposure studies (Wang et al. 2005). Recent trend in bioaccumulation and BCF research specifically has seen a rise in the use of radio labeled toxicants (Fisher et al. 1996). With radiolabeling, uptake and depuration kinetics as well as routes of entry are determined in different body parts of organisms down to cell organelles. Whereas the Bioconcentration factor determines uptake from water, the bioaccumulation takes into consideration both the water and the diet. In addition to routes of entry, bioavailability is also better determined using this technique. Rates and routes of metal bioaccumulation assist in the interpretation of accumulated body metal/toxicant concentrations in aquatic animals during their accumulation, sequestration, distribution and elimination in different aquatic species to explain how do aquatic organisms physiologically handle contamination (Wang et al. 2008). It was one of the aims of this study to use radio labelled tracers except that the infrastructure of the site, where the exposure experiments took place, could not sustain such trials.
82 Chapter 3.Bioaccumulation
Figure 8. a- sea urchins in stress, losing spines, exhibiting segregate positioning and not clinging to any surface, b- much healthier sea urchins with spines, clinging to wall of aquaria and aggregated in patches.
83 Chapter 3.Bioaccumulation
Compounds retrieved from the Gas Chromatography Mass spectrometry
During the GC-MS identification scan of both trials and for both treatments, aimed simply to confirm the presence of the organic compounds of interest, I came across few compounds with hydroxyl and methoxy prefixes. These compounds are believed to be metabolites produced by the cells of the roe. In table 8, the compounds of the treatment are listed, first according to their respective retention time, then according to their name as given by the Wiley Library to identify these compounds (relatively to the built in standards with a percentage of ion qualifier). I have taken the ions qualifier, which was higher than 70. Not all of these compounds have a chemical abstract number (ACS number), while some were not well referenced in the literature (their metabolic pathways, etc.). The underlined compounds were the few found in the control. It should be noted that it is not the aim of this study to investigate the metabolic pathways, the role or function of these substances, or to look for biomarkers (antioxidant, etc.). Although biomarkers can indicate exposure to xenobiotics but biomarkers rarely indicate the extent of exposure. I simply highlighted, where I could, the possible role of these compounds as extracted from the relevant literature (mentioned in the source‘s column). Some of the most interesting compounds were the 1-(methylthio)-9,10-anthraquinone believed to be an oxidation product of anthracene and squalene believed to be an antioxidant (reactive oxygen species ROS). It is recommended to monitor these compounds through administration of single toxicants and follow the change between outcomes of both control and treatment to pin down which toxicants produced which metabolites.
84 Chapter 3.Bioaccumulation Table 8. Compounds (some of them believed to be metabolites), retrieved from the scan mode using GC-MS, of the Paracentrotus lividus after bioaccumulation in both trials for both treatments. RT is the retention time on the Gas Chromatography-Mass Spectrometry. Each compound was compared to the Wiley 375 library along with the Chemical Abstract number (CAS), the qualifier of the masses in percent. The underline, bold and italic compounds were commonly found in control and spiked. It is believed that some of these metabolites have immunological functions. N/A (not applicable).
RT Library/ID Wiley 375 Gas Chromatography-Mass Spectrometry CAS Qualifier Suspected Role/Function Source 7.95 4-methoxy-4-methyl-2-Pentanone 000107-70-0 74 (Silvestre et al. 12.52 Seudenone 001193-18-6 86 Sex pheronome 2000) 14.32 Phorone 000504-20-1 87 Synthesis of acetone
14.99 .ALPHA.-ISOPHORONE N/A 90 intermediate in organic synthesis
18.80 5,6-Dimethoxy-1H-indazole N/A 80
21.48 Isoxylitone B 016695-73-1 97
24.72 3-(1',2'-Epoxy-2'-methyl-1'-propyl)-5,5-dimethyl-2-cyclohexen-1-one 077822-59-4 96
25.07 3,4,7-Trimethyl-1-indanone 035322-84-0 97
25.53 (11S)-12-Hydroxy-7.alpha.-eudesm-4-en-6-one N/A 83
27.16 (3S,5R)-3-(hydroxymethyl)-1,4,4,5-tetramethylcyclopentene 104086-71-7 72 Synthesized from Campholene
27.59 2-Methoxy-3-methyl-5-(1-pyrrolidinyl)-1,4-benzoquinone 077357-35-8 78
29.00 Valerenal 004176-16-3 78
29.76 Podocephalol 066656-01-7 90 (Kocsis et al. 30.40 6-Acetyl-7-Hydroxy-2,2-DimethylBenzopyran 019013-03-7 91 Antimicrobial 2002) 31.33 7-Isopropenyl-4,4,10.beta.-trimethyl-1,2,3,4,7,8,9,10-octahydronaphthalene 124666-93-9 90
34.25 8-amino-5-hydroxy-1-methylbenz[f]indazole-4,9-dione 107750-68-5 78
35.10 19-norkaur-16-ene deriv. 1H-2,10a-Ethanophenanthrene 076235-86-4 90
35.96 1-(4-methoxyphenyl)-3,3,5,5,6-pentamethyl-1-cyclohexene 098386-62-0 83
36.88 3,4-dihydro-7,12-dimethylbenz[a]anthracene N/A 90 a 37.58 4H-Furo[3,2-c][1]benzopyran-4-one, 2,3-dihydro-2,2,3-trimethyl-, (.+-.)- 116864-07-4 83 anti-inflammatory or immunomodulating action US Patent 41.16 1-(methylthio)-9,10-anthraquinone 002687-50-5 72 oxidation of Anthracene
85 Chapter 3.Bioaccumulation
Table 8. continued
RT Library/ID Wiley 375 Gas Chromatography-Mass Spectrometry CAS Qualifier Suspected Role/Function Source 41.30 2-methoxy-10-oxo-5,11a-dimethyl-7,8-dihydrophenanthrene 082612-84-8 80 1-(4-methoxyphenyl)-4-[[(4-methoxyphenyl)imino]methyl]-3,3-dimethyl-, (.+-.)-2- 44.31 Azetidinone 124870-65-1 83
45.30 6,11-Dihydroxy-3,10-dimethoxy-12H-benzo[b]xanthen-12-one 089140-95-4 80
48.29 5-acetyl-2-dihydro-6-methyl-2-oxo-4-phenyl-3-pyridinecarbonitrile N/A 83 b 52.92 Phorbol 017673-25-5 89 esters of phorbol are Tumor promoter
53.30 N-methyl-N-[4-[4-methoxy-1-hexahydropyridyl]-2-butynyl]-Acetamide N/A 95
53.80 PHENANTHREN-1-CARBOXYLIC ACID, 1,2,3,4,4A,9 001235-74-1 91 (Owen et al. 2004, Tikekar 60.22 Squalene 007683-64-9 72 Antioxidants, reactive oxygen species ROS et al. 2008)
65.38 8-hydroxy-3-[[1',5'-dimethyl)hexyl]3a,5b-dimethyl-3a,4,5,5b,6,7,8,9-octahydroindano[6,7-a]naphthalene N/A 97
68.56 4-Dehydroxy-N-(4,5-methylenedioxy-2-nitrobenzylidene)tyramine N/A 70
a) United States Patent 4900727 Kattige, Samba L. (Bombay, IN) et al. b) Through activation of protein kinase
86 Chapter 3.Bioaccumulation
Conclusion:
This study has set the BCF for anthracene, phenanthrene, pyrene, DDT/DDE, dieldrin, cadmium and lead in the sea urchin P. lividus and reaffirmed the BCF found in other studies for cadmium. It also highlights the effectiveness of sea urchin P. lividus as a good bioaccumulator to be used for risk assessment for the three categories of toxicants.
Sea urchin uptake of metal is the highest in roe followed in decreasing order by then test, spines and teeth with lower spike concentrations accumulating more than higher spikes due to a probable saturation of binding sites or as an immune response to toxicants. Accordingly, BCF studies for species should be conducted at more than one level as the ratio can decrease or increase according to toxicants‘ concentration but still BCF can help set thresholds for toxicity. This also highlights a potential problem for using BCFs as an indicator of toxicity if toxicity itself can change the BCF value.
Differences in lipid content between organisms will result in differences in bioconcentration factors for organic compounds. PAHs and OCPs have accumulated in the roe to similar ranges found in the literature for benthic organisms. The cocktail of toxicants might have acted in an additive way (combination of cadmium and 4,4‘DDE weakening calcites) or antagonistic way (organic competing with metals on binding sites). So the cocktail of toxicants in this case has helped accounting for the effects of this mixture (antagonistic, additive etc) without necessarily exploring the effects individually. It is recommended that some experiments be conducted for example to test any additive effect between DDT and cadmium. Finally, it must be noted that Algae and coelomic fluid have also taken up metals and PAHs and the toxicants in the algae might have biomagnified in urchins.
The prediction of chemical bioaccumulation by sea urchin is a key component required for the evaluation of both human health (as humans share few metabolic immune responses with echinoderms) and ecological risk assessments specifically in Lebanon, and more generally, in other parts of the world. Bioaccumulation can differ from region to region, even with organisms of the same species. Regulatory bodies and monitors could rely on such bioaccumulation models for establishing quality criteria for human health, and furthermore to set safe levels for sea food consumption, especially when more than one toxicant is present in consumed food.
87
Chapter 4
4.Depuration of Cd, Pb, Ni, V , dieldrin and 4,4’DDE from the sea urchin Paracentrotus lividus in the laboratory and in the field.
Abstract
Paracentrotus lividus was tested for its ability to depurate toxicants from its roe either naturally (field translocation) or in the laboratory (no free flowing water and assisted by EDTA chelation). Sea urchins were exposed to OCPs and metals for two weeks in the laboratory to stimulate bioaccumulation. Contaminants in the roe were measured and the urchins were left to depurate either in cages in a clean field location or in tanks in the laboratory. Tissues were sampled after days 1, 3, 7 and 14 of depuration. Depuration kinetics and biological half-lives were calculated for all toxicants used. Depuration in the laboratory was marked with fluctuations in the toxicant concentrations, which may be due to some mobilisation from other body parts such as the coelomic fluid. The chelating agent did not have noticeable advantage over natural depuration when used in very low concentration. In the field depuration study, all metals except cadmium depurated from the roe with a short biological half life (BHL) of 1.59 d-1 for vanadium, 2.36 d-1 for nickel, and 6.24 d-1 for lead, calculated after 24 hours of depuration. The OCPs, 4,4‘DDE and Dieldrin, depurated with a BHL of 1.34 d-1. Field depuration was more rapid and consistent than laboratory depuration. Depuration rates along with biological half-life indices could be used by the aquaculture industry to meet food safety guidelines.
88
Chapter 4.Depuration
Introduction
Toxicants accumulated by some marine biota pose a threat to both the organism and seafood consumers. Due to global trends of rapid economic development, population expansion and urbanization, the discharge of contaminants into the environment is increasing and with it, the probability of the biomagnification of toxicants, throughout the food chain. Some contaminants known to biomagnify such as persistent Organochlorinated Pesticides (OCPs) (Kannan et al. 2004), are currently banned worldwide, but are still produced by certain developing countries (Devanathan et al. 2009). The release of other toxicants, such as metals, may be more stringently regulated but these metals are still frequently found in elevated concentrations within the tissues of marine biota. This is of concern as there is great potential for consumption of tainted food to become a human health issue if an organism is subject to harvesting (WHO 2002). If environmental contamination cannot be prevented or removed, then the depuration of toxicants from seafood is a potential way to prevent human exposure.
Depuration is the process by which organisms are held in tanks of clean seawater under conditions which result in the expulsion of contaminants from different body parts of an organism, especially the intestinal contents, which prevents recontamination of the organisms via food that they have consumed (Lee et al. 2008). At the end of the nineteenth century and beginning of the twentieth century, typhoid outbreaks caused the spread of illness and death in many European countries and in the United States of America. Depuration was originally developed as one of several other means to address the outbreaks of typhoid associated with a large number of shellfish (caused by the bacterium Salmonella typhi) (Love et al. 2010). Depuration is currently used, not only to depurate bacterial loads, but also to monitor the elimination rates of many chemical contaminants in marine organisms.
The study of depuration and its kinetics is necessary for many reasons. If the organism is used as a bioaccumulator, it is imperative to study the depuration rate of that organism to determine whether it is fit for short or long-term biomonitoring programs. In many places, the exposure of aquatic organisms to environmental pollutants can occur through repeated pulses which can also fluctuate in concentration (Ashauer et al. 2006). In this case, depuration kinetics is needed for biomonitoring interpretation. Depuration studies are also needed to monitor and determine the length of time
89 Chapter 4.Depuration necessary for an organism to depurate a certain toxicant before it can be declared safe for human consumption. Finally, depuration of toxicants by marine organisms, is considered a survival tool to avoid toxicity. The steady state concentration reached following depuration of toxicants is therefore worth investigating.
Depuration kinetics can be calculated, for each toxicant, in body parts of the organism, by using 2 components exponential equation (Hubbell et al. 1965, Barron et al. 1990). Along with the Biological Half-Lives (BHL), it is of particular interest to the aquaculture industry. The BHL is the time required for half the quantity of a substance deposited in a living organism to be metabolized or eliminated by normal biological processes (Barron et al. 1990). It helps both researchers and the aquaculture industry predict the amount of time required for depuration to reduce toxicant levels to acceptable levels in the organism of interest. While researchers have the capacity to measure bioaccumulation and hence depuration in the laboratory at many time points, the aquaculture industry may not have such access and means and will rely on the BHL to estimate the amount of time a population of organisms inhabiting a certain area would need to depurate to safety levels.
Depuration can occur in an organism‘s natural environment as well as under controlled experimental conditions in the laboratory. Both approaches are used as depurating procedures in many aquaculture industries such as shellfish farming (Ho et al. 2000, Lee et al. 2008). In the field, the relocation of an organism (translocation), from a contaminated site to a cleaner site allows for the monitoring of depuration over a period of time under fluctuating environmental conditions. This is generally referred to as natural depuration (Ho et al. 2000). In the laboratory, depuration occurs in controlled conditions which generally do not fluctuate. Experimental studies may also be assisted by the addition of a chelating agent such as ethylenediaminetetraacetic acid (EDTA) known to complex with metals. The EDTA molecule can bind to metal ions by forming tightly bound complexes which allows a more rapid removal of metals from the organism (Gil et al. 2011). In the aquaculture industry depuration rates may be increased by the provision of clean food or filtered water (Little et al. 2003).
Depuration for the purposes of preparing seafood must involve the removal of contaminants from tissues that are consumed to below regulatory levels. This is not the same as detoxification which may render contaminants non-toxic but allow them to remain in the tissue. At the cellular level, detoxification can include phagocytosis, binding to metallothionein, production of Reactive Oxygen Species or cellular efflux through binding to cell organelles (Chia et al. 1996, McGeer et al. 2003, Vijver et al. 2004, Newman et al. 2008, Wang et al. 2008). Detoxification in this sense,
90 Chapter 4.Depuration can be temporary or permanent pending on the nature of the toxicants and the health of the organism. Depuration, however, is a permanent process and cannot be reversed unless further contaminant exposures occur.
Two common subclasses of contaminants that may need to be depurated from seafood are OCPs and metals. OCPs, such as DDT and dieldrin, have been found in marine biota and are at levels that cause disruption to endocrine systems (ATSDR 2002b). They exhibit high octanol/water partition coefficients (Log Kow>4) and are selectively partitioned from water into the lipid tissue of aquatic biota (Elia et al. 2007). It is therefore predicted that the highest organic compounds concentrations are present in high-fat tissues such as milk glands in mammalians and the roe of sea urchins. Some metals can also be toxic to organisms, including humans, affecting nervous and reproductive systems even at low concentrations, with sufficient evidence for carcinogenicity in humans (ATSDR 2007, 2008). Metal toxicants such as cadmium and lead are ubiquitous and are neither biodegradeable nor can be eliminated by incineration processes (Accornero et al. 2004). Because metals persist in the environment, they have been used as tracers of oil exposure. In marine organisms, metals such as nickel and vanadium have been used as tracers of exposure to oil deposits along the coast (Amiard et al. 2004). Depuration models would predict that most metals would be depurated by marine organisms using existing cellular detoxification mechanisms while the depuration of OCPs would be limited to the reduction in lipid content of the organism (Newman et al. 2008).
On the eastern side of the Mediterranean Sea, toxicants emerging from untreated industrial, agricultural and domestic waste are discharged directly into the sea and are taken up by sea urchins as illustrated in Chapter 2. Off the Lebanese coast, the sea urchin Paracentrotus lividus, is consumed directly after harvesting and prior to any depuration process. It is an important resource for fisheries due to its commercial value (Guidetti et al. 2003b).The lack of depuration prior to consumption may have consequences to the human health, as little is known regarding the uptake and depuration capacity of this organism. One study by Warnau (1995b) examined the depuration rate of cadmium in P. lividus, but to the best of our knowledge the depuration of other metals or OCPs have not been studied in this species.
Aims The aim of this study was to examine depuration of cadmium, lead, nickel, vanadium, dieldrin and DDT from the sea urchin P. lividus under conditions of natural depuration by translocation and under laboratory assisted conditions with the addition of the chelating agent EDTA. Emphasis is placed on depuration from the roe as this is the edible part of the sea urchin and because it is a
91 Chapter 4.Depuration common site where both OCPs and metals accumulate. Results of this study can provide sea urchin fishery managers and the aquaculture industry with tools to help them sustain fisheries and the trade of seafood produced, whilst complying to food safety limit methods.
Method
Collection of Sea Urchins
Sea urchins species P. lividus, were collected in May 2007 by SCUBA diving between 5 and 10 m depth off the coast of ―Tripoli‖ (34o 29‘ 21.74‖ N and 35o 46‘ 27.33‖ North of Lebanon). Sea urchins were transported to the Marine Laboratory at the American University of Beirut (Biology Department) in clean containers of aerated seawater. Prior to experimentation, The sea urchins (260 in total, average size ca 4cm) were acclimatized to laboratory conditions for one week in eight glass aquaria tanks (60L) filled with natural sea water constantly aerated (Salinity: 38-41%o, temperature: 20.0±3°C, 12/12 h dark/light cycle). During the acclimatization period, the tanks were monitored daily and dead specimens were removed. The sea urchins were distributed randomly and evenly between the tanks (ca 30 sea urchins/tank).
After one week, 12 sea urchins in total were randomly chosen from the eight tanks in order to determine the bioaccumulation baseline prior to the addition of toxicants (Sea urchins from this collection were labeled ‗baseline‘). The experiment was run for a total of five weeks. This included a one week acclimatization period, two weeks of spiking and two weeks of depuration either in the laboratory or the field (see details below and Table 1). A depuration period of two weeks was chosen as it has frequently been used in the literature and because it is an economically viable amount of time to be used in the aquaculture industry. This period also minimized the risk of cages being tampered with in the field. Depuration in the laboratory is hereafter referred to as laboratory depuration while depuration in the field is referred to as field or translocation depuration in reference to the location where the depuration experiments took place (Table 1).
Experimental set-up and Spiking (Exposure)
Eight tanks were assigned to the experimental treatments (Control and toxicant feed) with 4 replicate tanks per treatment. The toxicant spiked treatment consisted of four 60 litre-tanks each
92 Chapter 4.Depuration spiked with a 50 ml solution (deionized water: acetone 50:50) containing a cocktail of toxicants and achieving final tank concentrations of 50µg/l (50 ppb) for each of the dieldrin and 4,4‘DDT and 100 µg/l (100 ppb) for nickel, lead, cadmium and vanadium. The control treatment consisted of four 60 litre-tanks to which a 50 ml solution (deionized water: acetone 50:50,) that was not spiked with toxicants, was added. The choice of toxicants was based on knowledge gained in chapter 2 regarding bioaccumulation and the quantity used was considered sufficient to raise tissue concentrations to a level worth following during depuration. Organic toxicants were diluted in their proper dissolving solvents before being added to the carrier solvent (acetone). Inorganic metal toxicants (cadmium, lead, nickel and vanadium) were diluted in deionized water (acidified with HCl) before being added to the acetone fraction.
The choice of the first two metals; cadmium and lead was based on their toxic effect on the reproduction of sea urchin, the health of consumers and because they are by-products of industrial activities (Au et al. 2003, Xu et al. 2010). Vanadium and nickel were chosen in light of a recent oil spill in Lebanon (Coppini et al. 2010) as these two metals are usually found in the dispersed fuel oil (Tronczynski et al. 2004). Dieldrin and 4,4‘DDE were chosen because they are still found in the tissue of terrestrial and marine organisms with devastating consequences (Scanes 1997, Taniguchi et al. 2009).
Immediately after preparation of the tanks with the spiking solution, the contents of each of the tanks were mixed by slightly increasing the air bubbling. Contaminant exposure in the laboratory was maintained for two weeks with water changed after one week and spikes renewed at this point. 24 hours prior to the water change, the sea urchins were fed algae (Ulva species) to the equivalent of approximately 2% of their body weights. Urchins were again fed 24 h prior to the start of the depuration period. The feeding schedule was organized in order to minimise the amount of toxicant absorbed by the food which could result in reduced toxicant levels in the water of the tanks. Faeces and shed spines were removed daily during this period to avoid uptake of toxicants into these waste products.
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Table 1. Laboratory and translocation experiments laid down in a timetable. The concentrations for the baseline and the concentrations after feed are commun for both experiments. Laboratory and translocation experiments can be referred to by the name of the location where experiment took place. Locations of On arrival Baseline First week Second week 1st day of 3rd day of 7th day of 14th day of experiments determination depuration depuration depuration depuration Laboratory and Acclimatisation Sea urchins Toxicant Toxicant See individual See individual See individual See individual Field Removal of dead removed and mixture mixture experiments experiments experiments experiments Translocation individuals frozen to introduced on introduced on below below below below Monitoring determine the first day the last day of Time since Cumulative Cumulative Cumulative time physical and baseline toxicant sea urchin fed the previous arrival of sea time since time since since arrival of chemical concentrations on the 6th day week, urchins: 3 arrival of sea arrival of sea sea urchins: 5 parameters of baseline is water changed sea urchins fed weeks and 1 day urchins: 3 urchins: 4 weeks water common for both on the 7th day. on the 6th day of weeks and 3 weeks Both experiments Food : ad libitum experiments Time since the second week days ended at the 7th Duration: 1week Duration: 1 hour arrival of sea the water was day of this week. before spiking urchin: 2 changed on the Cumulative time weeks. 7th day for experiment Individuals since arrival of were sampled to urchins: 1 week quantify their and 1 day bioaccumulation sea urchins were split between two experiments. No more toxicants added after the change of water Time since arrival of urchins : 3 weeks
94 Chapter 4.Depuration
Table 1. continued
Locations of On arrival Baseline First week Second week 1st day of 3rd day of 7th day of 14th day of experiments determination depuration depuration depuration depuration Laboratory Half of the sea 24 hours after 72 hours after 7 days after 14 days after urchins remain addition of addition of addition of addition of in tanks, EDTA EDTA, EDTA, EDTA, EDTA, was added 3 sea urchins 3sea urchins 3 sea urchins 3sea urchins collected collected collected collected referred to as 1 reffered to as referred to as referred to as day after 3 days after 7days after 14days after depuration, depuration. depuration. depuration. water changed, Water Sea urchin fed Sea urchin fed on EDTA added changed and on the sixth the sixth day EDTA spiked day , EDTA end of experiment added again on the 7th day of this week Field The rest of the Diving and Diving and Diving and Diving and Translocation sea urchins were sampling from sampling from sampling from sampling from caged and taken cages cages cage cages in the field for feeding .urchins fed on the depuration animals on the 6th day 6th day end of experiment on the 7th day
95 Chapter 4.Depuration
Laboratory Depuration At the end of the toxicant spiking period, a subset of sea urchins was sampled (3 sea urchins from each tank) to determine the immediate bioaccumulation concentrations. Following this, half of the sea urchins were randomly chosen from the tanks and placed in cages (figure 1) to be moved to the field (12 to 15 sea urchins were placed in each cage). The remaining urchins were used for the assisted lab depuration experiment (details of field translocation below). The water was then changed in the tanks at least three times, faeces were removed and aquaria walls scrubbed with a clean sponge to wash off all residues of toxicants before the start of the lab depuration. Water was then refilled for the fourth time and EDTA 1 mg l-1 was added to all the tanks to start the depuration experiment. At days 1, 3, 7 and 14, 3 sea urchins were removed from each tank and prepared for chemical analysis. Six days into depuration, the remaining urchins were fed, water was changed, EDTA added and on the seventh day, algae were removed. Water was changed daily till day 13, when the remaining individuals were fed once more. Water was changed and spiked with EDTA. This was the end of the laboratory depuration experiment. The tanks were monitored every other day for parameters such as oxygen, salinity, temperature and pH.
Translocation depuration
In order to assess the potential for depuration of toxicants (Organic and inorganic compounds) under field conditions, a translocation experiment was conducted in Batroun Bay, North of Lebanon. Sea urchins were transferred from the laboratory where they had been exposed to toxicants (as above) to Batroun Bay (South of Tripoli) (34o 15‘ 20.94‖N and 35o 39‘ 27.04‖) at the Center for Marine Research, an area relatively free of sewage, industrial and agricultural activities (GreenPeace et al. 1997). Each cage was labelled and was placed in a clean sea water bucket equipped with air filters working on batteries during transportation. At the field site, cages were stationed below lowest tide protected by rocks at 3-5 meter depth to eliminate exposure to stressful events. Cages were bundled together with ropes (0.5 m distance between each two) secured by weights. With these precautions, the majority of sea urchins survived throughout the whole translocation experiment. However, the final few urchins in the control treatment cages were emptied by poachers 3 days before the end of the translocation experiment, resulting in a fewer data for the final census. Similar to the laboratory study, urchins were sub-sampled from the field translocation experiment on the same days as the assisted lab depuration experiment (days 1,3,7 and 14 after depuration).
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a
b
Figure 7. a-Aquaria at the Marine Laboratory of the American University of Beirut with sea urchin before removal of faeces and after feeding. b- cages labelled and used for translocation experiment in which sea urchins are put for a period of two weeks. All cages were bundled together with one rope to keep them in one spot and anchored.
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Analytical procedures
Sample preparation At each subsampling time, three sea urchins were randomly removed from each tank and were rinsed with 1 % nitric acid and acetone before being further processed. The sampled sea urchins were then measured for diameter size, weighed and other body characteristics (e.g. roe weight and color were noted). The sea urchins were excised and separated into the following body parts: roe, spines, teeth and test. The roe was extracted using stainless steel or polyvinyl chloride (PVC) utensils. All samples collected from the same tank within each treatment group were pooled. The roe was freeze-dried, homogenized then used to determine the organic and inorganic loads (details below). All dissections and analyses of samples were undertaken at the ISO 17025 accredited environmental core laboratory at the American University of Beirut.
Metal analysis The metal analysis was done according to the protocol highlighted in the temporal variation chapter.
Metal Quality Control The quality control protocol used for the metal analysis is the same as the protocol described again in the temporal variation chapter. However, in this chapter, data was further treated for lead and cadmium in the roe (50% of data) when lead levels for the CRM was outside limits highlighted in the certificate. Results were expressed as microgram of metal per g of dry weight (µg g-1). For statistical purposes values less than the detection limit were reported as half the limit of quantification. The limits of quantification, for all metals, were 1 µg l-1 and 0.01 µg g-1 in water and tissues respectively. All methods used for digestion and analyses were from internationally approved methods with some modifications (EPA-USA-200.8 1994, EPA-USA-3052 1996, APHA 1999).
Organic method: OCP extraction, clean up, analysis and confirmation of OCP Sample extractions were performed using a Microwave Extraction System (Milestone), saponified and cleaned up then inject on a gas chromatography electron capture detector (GC-ECD) and on a gas chromatography- mass spectrometer (GC-MS) using scan and selective ion monitor (SIM) modes for confirmation as detailed in the bioaccumulation chapter.
OCP Quality Control To evaluate both the extraction efficiency for the target compounds and the efficiency of GC–MS for analysis of these compounds, recovery studies were carried out using both a standard addition and an internal standard. The OCPs in the samples were identified by a combination of a retention time match and mass spectra match against the external calibration standards. Quantification was
98 Chapter 4.Depuration performed by the method of internal standardization using Endosulfan II as an IS. External standards of OCP were used to calculate response factors by comparing their peak areas with the IS. The limits of quatification were 0.02 µg l-1 and 0.05 µg g-1 in water and tissues respectively. Analytes‘ concentrations were corrected for extraction efficiency by multiplying the average recovery of the surrogate/internal standards by each analyte concentration detected in the sample.
Chemstation software (B.01.03 /SR1) was used for data acquisition. During the analysis, quality control (QC) was carried out on each set of samples to check performance of digestion (blank of digestion, spiked blanks, duplicates, sample spiked) and to check performance of instrument (tuning, ion ratios, standard curve, continuous calibration standard) with each batch of solvents used before and after maintenance of column and/or instrument consumables (washers, septum etc.).
Depuration kinetic rate determination
Toxicokinetic- toxicodynamic (TK-TD) models simulate the time-course of uptake, biotransformation, and elimination of toxicants in the organism (toxicokinetics, TK), which may then determine the time-course of damage and organism recovery as the prerequisite for toxic effects (toxicodynamics,TD) (Ashauer et al. 2007). Depuration rate constants for the individual compounds were evaluated first using a simple first-order one-compartment model (Spacie and Hamelink, 1982) if the depuration follows a linear regression mode. Under steady-state conditions, the equation describing the kinetic depuration can be written as it was used in literature by many authors (Choi et al. 2003, Gatidou et al. 2010, Muscatello et al. 2010). So using depuration kinetic equations, purging rates of the sea urchins were to be calculated at steady state by the equation: -kt -1 Tt=T0e , where t is the depuration period (day), Tt is the total toxicant amount (µg g ) in sea -1 urchin roe at time t (concentration after x day of depuration), T0 is total toxicant amount (µg g ) in sea urchin roe at time 0 (concentration after two weeks of toxicant addition) and k is the depuration rate constant. The depuration rate constants were calculated from the slope of the linear regression obtained by plotting LogeTt against t. If the linearity is not observed with or without data log transformation, then the kinetics were fitted by a 2-component exponential equation following the method of Hubbell et al.(1965). Separate exponential fits were used for the first day (just two points characterizing the fast/short depuration process referred to in the equation as s) and the subsequent slower purging (long term depuration referred to as l).
-kst -klt -1 Tt= T0se +T0le , where t is the depuration period (day), Tt is the total toxicant amount (µg g ) in sea urchin roe at time t (concentration after x day of depuration), T0s is the short lived component
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(i.e. the loss of toxicants that are poorly associated within the organism, Tol is the long lived component that is tightly bound in the organism (total toxicant amount (µg g-1) in sea urchin roe at (concentration after two weeks of toxicant addition) and k is the depuration rate constant.
The depuration rate constants for both the one phase and two phases depuration (from day 0 to day 1) and the consequent rates are tabulated (Table 4). The biological half-lives of toxicants in the sea urchin i.e. the time required to reduce by one-half, the toxicants concentrations in P. lividus roe tissue (t 1/2) were determined solely from the k of depuration and were calculated using the following equation
-1 -1 t 1/2 = (t 1/2 = ln 0:5 k .= 0.693 k ) for the one component.
For the two components model, a biological half life may be calculated for each component (Tb1/2s and T Tb1/2l). Fitting of the data to each of the two exponential components was assessed on semi
-1 -1 log transformed data using the linearity test. t 1/2 = (t 1/2 = ln 2 k .= 0.693 k ) for the two
components.
Data treatment and statistical analyses A one factor ANOVA (α=0.95) was used for each of the four weeks separately to check for differences in the size, total and roe weight of the sea urchins with treatments (control, and toxicants spiked) as factor. A normality test (Q-Q plot) and a homogeneity test (Levene) were checked before the use of ANOVA (Quinn et al. 2002). Differences in sizes or in roe weight may have meant that these two parameters should be considered as cofactors in any ANOVA analysis on the bioaccumulation and depuration. I was planning on using a repeated measures ANOVA (significance at 0.05) to detect differences in depuration of sea urchins with days of depuration as ―within factor‖ and depuration location (laboratory depuration v field translocation depuration) and treatments (control and spiked) ―between factors‖ in the roe. The assumption of normality was not met , analysis showed the data skewed to the left (because it was combining both accumulation and depuration), and since no known transformation would have rendered this normal, a permutational analysis (permutation of variance PERMANOVA) that does not assume homogeneity of variances or normality of distribution was used instead. Our results from the PERMANOVA (using PRIMER v6) (Clarke et al. 2006)) are tabulated in the results section. The post hoc for within days is interpreted from the graph. SPSS software (version 18) (2010) and Primer (v.6) were used for all statistical analyses. When running the statistical analysis of Repeated ANOVA on roe contamination, and so that I do not exclude the last week of the translocation experiment because
100 Chapter 4.Depuration of the theft of urchins cages mentioned previously, the values of the control of that week (2nd week of depuration) were filled with the values obtained by the same controls but after one week of depuration. Care was taken in the interpretation of the interaction between days and experiments for the second week due to this assumption of data translocation.
Results
Size, Total Weight and Roe Weight During the four weeks of the two experiments (depuration in laboratory and depuration via translocation), the sea urchins were of similar size (3.6 to 4 cm), roe weight and total weight which eliminates these parameters as cofactors affecting bioaccumulation and depuration (Figure 2) or inciting differences in depuration rate of both experiments (translocation and laboratory depuration). The average lipid percentage of 12 sea urchins randomly sampled from both experiments was 14% in roe dry weight.
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Figure 8. a-Size in cm (Mean±SE), b-total weight and c-roe weight both in g (Mean±SE) of the sea urchin Paracentrotus lividus used for both the depuration experiments (the laboratory assisted depuration experiment and the field translocation experiment) and for both treatments (control and contaminants). The control for the field translocation experiment ―14 days after depuration‖ is missing due to theft. Before and after toxicant addition are common to both experiments.
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Laboratory and field depuration for cadmium, lead, nickel, vanadium Depuration curves for the sea urchin P. lividus showed differences in elimination rate between the field and the laboratory with some fluctuations seen mostly in the latter. The shapes of the curves indicate that a steady state was still not reached by day 14; with some exceptions. Lead accumulated in the roe after spiking and began decreasing after the first day of depuration in both experiments (Laboratory and translocation) (Figure 3). In the laboratory experiment, the toxicant treatment dropped by three quarters and reached the same level as the control by day 14 (from 1. 03 to 0.28 µg g-1). In the field translocation, the same pattern as the laboratory depuration was initially observed, however; a slight increase in lead concentrations was detected at the end of the experiment. The control was steady in both experiments. There was no interaction between depuration location and treatment (table 2).
Results for Cadmium in the laboratory were different from results obtained in the field (Table 2). In the field translocation it depurated for the first three days then after the third day concentrations began to increase and reached levels higher than before translocation but this pattern did not result in a significant difference over the test period. As for the laboratory experiment, no depuration was detected in the contaminant treatment. The control in both experiments (laboratory and field) behaved similarly with concentrations remaining unchanged throughout the 2 weeks (Table 2, Figure 4).
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Figure 9. Bioaccumulation (mean ±SE) in µg g-1 of lead in the roe of sea urchin Paracentrotus lividus for contaminant treatments (control and toxicant) before and after the addition of lead and then after one, three, seven and fourteen days for both depuration experiments (translocation in the field and laboratory assisted using EDTA).
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Figure 4. Bioaccumulation (mean ±SE) in µg g-1 of cadmium in the roe of sea urchin Paracentrotus lividus for contaminant treatments (control and toxicant) before and after the addition of cadmium and then after one, three, seven and fourteen days (mean ±SE) in µg g-1 for both depuration experiments (translocation in the field and laboratory assisted using EDTA).
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Table 2. Repeated Measure ANOVA using PERMANOVA (999 permutations), testing for differences in bioaccumulation of lead, cadmium, nickel and vanadium in the roe of the sea urchin Paracentrotus lividus with days of depuration as factor within subjects and contaminant treatments and location of depuration experiments(field and laboratory) as factors between subjects at a significance of α= 0.05. Data was not transformed.
Lead Cadmium
Source df MS P(perm) df MS P(perm)
Location of depuration experiments 1 0.011727 0.721 1 4.258 0.008
Contaminant treatments 1 3.7914 0.001 1 112.59 0.001
Days 4 0.68689 0.001 4 1.2602 0.073
Locations x Treatments 1 0.11102 0.242 1 2.3458 0.048
Locations x Days 4 0.052824 0.654 4 1.6744 0.028
Treatments x Days 4 0.14725 0.156 4 1.1905 0.078
Locations xTreatments x Days 4 0.07228 0.498 4 0.88162 0.197
Res 60 0.083848 60 0.54858
Nickel Vanadium
Source df MS P(perm) df MS P(perm)
Location of depuration experiments 1 0.0008913 0.9 1 0.56451 0.395
Contaminant treatments 1 0.011132 0.67 1 0.48592 0.404
Days 4 0.10968 0.124 4 3.2758 0.005
Locations x Treatments 1 0.0001435 0.968 1 0.2566 0.513
Locations x Days 4 0.001781 0.999 4 0.99284 0.2
Treatments x Days 4 0.042279 0.56 4 1.0839 0.16
Locations xTreatments x Days 4 0.014589 0.929 4 0.36521 0.71
Res 60 0.058318 60 0.66639
Nickel in both experiments (laboratory and field) did not show any statistical differences nor did the different treatments in each of the experiments (Figure 5, Table 2).
For vanadium, depuration in both experiments were similar, after a slight increase at the end of the spiking period for both the control and contaminated treatment (Table 2) I observed a decrease in concentration with time at a rate depicted in Table 4 till it reached a steady state for both experiments and both treatments. The value of vanadium after 14 days of depuration was similar
106 Chapter 4.Depuration between the contaminated treatment and the control treatment (Figure 6, decrease of 50%). I also detected a fluctuation in the control in both experiments, which was more evident in the translocation experiment but was not significantly different (Table 2).
Figure 5. Bioaccumulation (mean ±SE) in µg g-1 of nickel in the roe of sea urchin Paracentrotus lividus for contaminant treatments (control and toxicant) before and after the addition of nickel and then after one, three, seven and fourteen days for both depuration experiments (translocation in the field and laboratory assisted using EDTA).
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Figure 6. Bioaccumulation (mean ±SE) in µg g-1 of vanadium in the roe of sea urchin Paracentrotus lividus for contaminant treatments (control and toxicant) before and after the addition of vanadium and then after one, three, seven and fourteen days for both depuration experiments (translocation in the field and laboratory assisted using EDTA).
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Laboratory and field depuration for OCPs 4,4’DDE and Dieldrin in roe OCPs depuration curves for the sea urchin P. lividus were characterized by a fast depuration phase on the first day followed by a slow phase for both pesticides 4,4‘DDE and dieldrin in the field translocation experiment. Still in the field translocation experiment, dieldrin concentrations returned nearly to baseline levels within 14 days (Figure 7 and Table 3). Depuration in the laboratory was characterized by a fluctuation in dieldrin concentrations after the major depuration step before settling down to a steady state whereas in 4,4‘DDE increased from 11 to 24 µg g-1 (Figure 8).
The control behaved similarly in both depuration locations except for a slight increase in day 1 in the dieldrin translocation and depuration, which although not significant statistically may be due to volatility of the pesticides from treatment tanks, their solubility in the air and the absorption of these pesticides by the water in the control tanks.
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Figure 7. Bioaccumulation (mean ±SE) in µg g-1 of dieldrin in the roe of sea urchin Paracentrotus lividus for contaminant treatments (control and toxicant) before and after the addition of dieldrin and then after one, three, seven and fourteen days for both depuration experiments (translocation in the field and laboratory assisted using EDTA).
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Figure 8. Bioaccumulation (mean ±SE) in µg g-1 of 4,4‘ DDE in the roe of sea urchin Paracentrotus lividus for contaminant treatments (control and toxicant) before and after the addition of 4,4‘ DDE and then after one, three, seven and fourteen days for both depuration experiments (translocation in the field and laboratory assisted using EDTA).
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Table 3. Repeated Measure ANOVA using PERMANOVA (999 permutations) testing for differences in P.lividus' bioaccumulation of dieldrin and 4,4‘ DDT in the roe of sea urchin Paracentrotus lividus with days of depuration as factor within subjects and contaminant treatments (control and toxicant) and location of depuration experiments(laboratory and field) as factors between subjects at a significance of α= 0.05. Data was not transformed.
Dieldrin 4,4' DDE
Source df MS P(perm) df MS P(perm)
Location of depuration experiments 1 158.75 0.376 1 158.75 0.385
Contaminant treatments 1 10923 0.001 1 10923 0.001
Days 4 1561.2 0.001 4 1561.2 0.001
Locations x Treatments 1 29.538 0.831 1 29.538 0.813
Locations x Days 4 94.299 0.814 4 94.299 0.797
Treatments x Days 4 1427 0.001 4 1427 0.001
Locations xTreatments x Days 4 177.64 0.482 4 177.64 0.491
Res 40 192.63 40 192.63
Depuration rate and Biological Half Life The depuration rate initially calculated for both locations of depuration (field and laboratory), for the contaminant treatment, showed a difference between the locations in the depuration kinetic rates for all toxicants except OCPs. However, the fluctuations in the laboratory experiment concentrations have dictated that the kinetic calculations will be taken, solely, from the field experiment except for nickel. The rates of field translocation are taken from the plotting of the loge
Tt are determined for the two phases of depuration (fast: first 24 hours and slower: days 3 to 14) and are presented in table 4. The fastest depuration rate was scored, in the field, by the OCPs(0.52 d-1) while the slowest was recorded by lead (0.111 d-1) (cadmium was excluded since it did not depurate). Also in the field, the highest Biological Half Life (BHL) was calculated in lead (6.244 days) while the lowest BHL was for the OCPs (1.36 d-1) (Table 4).
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Table 4.Depuration rate kinetic calculated for lead, cadmium, nickel, vanadium, dieldrin and4,4‘DDE in the roe of sea urchin Paracentrotus lividus for the field translocation experiment. First depuration constant (k s ±SE in d-1 short s) was calculated between the end of feeding toxicants and the first day/third day of depuration and is expressed as constant d-1, the second point (k l ±SE in d-1 long l)was taken between the first day/third day of depuration and the 14th day and is expressed as constant d-1. Biological Half Lives of toxicants (BHL) are expressed as constant d-1 for each component separately. R2 coefficient of correlation (for linear or exponential) as well as significant ―p‖ at 0.05 are tabulated along with the type of kinetic used (one linear component vs two components).
Type of V Ni Cd Pb 4,4‘DDE Dieldrin components
One or two Not kinetics 2 1 applicable 2 2 2
R2 0.3647 0.9766 0.903 0.979 0.945 p 0.005 0.9 0.001 0.001 0.001
0.435 0.293 0.111 0.519 0.508 k s (SE) in d-1 (0.05) (0.134) (0.024) (0.044) (0.021)
0.02 0.02 0.036 0.08 k l (SE) in d-1 (0.001) (0.011) (0.001) (0.002)
BHL t ½ s(days) 1.59 2.36 6.244 1.33 1.36
BHL t ½ l (days) 34.65 34.65 19.25 8.66
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Discussion
The depuration kinetics of toxicants measured varied greatly between the different chemicals that P. lividus were exposed to. Cadmium and vanadium were accumulated however, unlike vanadium, cadmium was not depurated. Nickel did not appear to accumulate within individuals when they were spiked and depuration of baseline levels found in organisms did not occur. Lead was the only metal in which both accumulation and depuration were observed. Perhaps it is due to the way lead is stored inside cell organelles which makes it easy for it to be eliminated. My results are in agreement with other published studies for benthic animals in more than one aspect. For example, in OCPs there was a sudden drop in concentration followed by a slow depuration phase (Choi et al. 2003). Another aspect in agreement with other published studies, is the fact that sometimes in the middle of the depuration phase a sudden increase in levels was observed followed again by a drop that eventually takes it to a steady state (Antunes et al. 2007b).
Depuration rate is affected by a) mode of exposure (chronic vs sudden), b) nature, characteristics, levels and classes of toxicants (high kow for organics, competition to ionic binding sites for metals, high levels vs low levels), c) physiological target/binding sites(incorporation of toxicants inside the cells like granules, ligands etc), d) cleanliness of seawater and e) the time needed for depuration.
The sudden drop in depuration can be explained by the sequestration method used by the organism in its defence and by the type of exposure. According to Gardinali (2004), depuration rate is a function of the type of exposure (chronic vs sudden). Elimination of all toxic compounds proceeds at a slower rate in chronically exposed populations in comparison to a newly contaminated one. In my experiments, if one considers the nominal value of the spike and the time of exposure (2 weeks) and the lower values in the field to which the organisms were exposed to before collection, it can be said that a sudden exposure has been mimicked in the laboratory.
Toxicity has also been noticed to affect depuration rate, in that the more toxic the organic compound, the longer depuration of the toxicant will take (e.g. Poly Chlorinated Biphenyls (PCBs) require from 42 to 60 days in oysters (Gardinali et al. 2004)). This is likely the scenario for metals; the more damage a toxicant does at the cellular level the less able the organism may be of up-regulating detoxifying mechanisms (Newman et al. 2008) and vice versa the longer the depuration, the higher the possible toxic effects. Even within the species itself, not only the mode of uptake and bioaccumulation will affect the
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Chapter 4.Depuration
depuration as discussed above (the chronic versus the sudden), but also high levels of toxicants versus low levels (where low levels are accumulated more than high levels, the latter causing saturation of binding sites). During sudden exposures some organisms sequester toxicants in granules for fast elimination (Wang et al. 2005, Rainbow 2007). This defensive technique is in contrast to incorporating them inside the cellular organelles (ligand module) that may be more common during chronic exposures (Veltman et al. 2010). In any event, even when contaminants are detoxified or sequestered in granules inside the cell and are probably inactive physiologically, their toxicity can still affect the organism if uptake of toxicants continues. For example, the organism may resort to sequestering these toxicants (by formation of granules), however, the building up of granules may physically affect elasticity of the cell (Buxboim et al. 2010) and therefore may cause the death of the cell by bursting especially if uptake continues and granules are not excreted.
Time could be an important factor influencing toxicant concentrations at the end of the depuration period. It could interact with the clean sea water to give the organism a chance to achieve a homeostasis. This experiment is time limited, and I cannot predict whether sea urchins would have been able to purge the cadmium from their roe if the depuration period had been longer. Warnau et al. (1997) noted a biological half-life of one year for cadmium. Hence, this aspect can be investigated in future studies, focusing on means of purging cadmium.
Metals The depuration rate of lead by P. lividus in the field on the first day was 0.111 day-1. In general I would predict faster rates of depuration in clean field environments due to the constant flushing of seawater and provision of clean food (Newman et al. 1998). The slight drop in lead concentration, first seen in the laboratory, might have been indicative of increased complexing through the addition of EDTA. However, EDTA, as used in this study at 1 mg/L in sea water, was probably immediately saturated by soluble Mg and Ca (concentration of Mg alone is ca 1.3 g/Kg SW and that of Ca 0.41 g/Kg SW). Even if the EDTA concentration was increased in my tanks it is believed that it would not have complexed with the cadmium as was noticed by Resgalla et al., (2012) when an attempt of detoxification was undertaken in seawater on A. lixula. So EDTA could not have been involved in any heavy metal depuration. In principle, EDTA has been approved and used by FDA as a chelation therapy to treat acute and chronic heavy metal poisoning by pulling toxins (including lead, cadmium, and mercury) from the bloodstream (Burns et al. 1995, Lin et al. 1999). The initial thought was that EDTA would have a dual role to play
115 Chapter 4.Depuration
in this study: It would sequester the metals purged by the organism rendering them inert and hence not bioavailable to be taken up again or it would complex with metals inside the roe tissues, speeding up the purging of toxicants. However, with the low EDTA concentration used it was not achievable. Increasing this concentration, nonetheless, has a downside to it as it can cause toxicity by binding to essential metals (Powell et al. 1999) which might lead to disruption in normal cellular process and level of cell dysfunction.
In the translocation experiment, the increase in lead and cadmium concentrations observed during the depuration period (day14 for lead, day7 for cadmium) may be a result of natural fluctuations between replicates or a result of remobilisation of the element from other body compartments (calcites, digestive tract or coelomic fluid). This process has been observed in previous studies in relation to organic compounds however the principle still applies (Choi et al. 2003, Antunes et al. 2007b). If, remobilisation is indeed involved, it is not clear why increased roe concentrations would not also be observed under laboratory depuration condition. Another possibility, to be considered in this case, is a potential sudden contamination being released in the translocation area. However, it is regrettable that the control was not available from the second week of the translocation experiment (due to poaching) to rule out the second possibility.
As for the BHL of lead in literature, it has been calculated to be around 30 to 45 h for organic lead in rainbow trout (ATSDR 2007). This figure is not far from our BHL for inorganic lead in the sea urchin, which was 144 h. Of course, si nce each species has its own mean of bioaccumulation and depuration mechanisms the BHL extracted from literature are just mentioned here to give an idea of the scale rather than for comparison.
It is not clear why depuration of cadmium was not observed in the laboratory study whilst the field translocation observed cadmium fluctuations. In the laboratory, uptake into the roe from coelomic fluid or the digestive system may have supplied the roe with more cadmium (Coelomic fluid, from the previous experiment, has been accumulating the toxicants), whereas urchins in the field translocation study may have been able to depurate the coelomic fluid and the digestive tract more rapidly due to higher rates of water flow. In general, this lack of depuration in the roe is in agreement with Warnau et al. (1995b) who also noted that the loss of cadmium was not significant if the depuration period was less than a year. The findings are also in agreement with studies reviewed by
116 Chapter 4.Depuration
Reboucas Do Amaaral et al.(2005). Those researchers noted that Cd was not depurated from the soft tissues of oysters. They observed that concentrations in the oysters‘ soft tissues were practically the same after 90 days of depuration following controlled exposures in the laboratory. Other field studies have shown that Mytilus californianus was able to release its cadmium significantly but slowly in just a few days (up to 30% in 2 days) if the Cd concentration in the water was low (1 µg/L). However, depuration was not observed if the Cd concentration was as high as 10 µg/L (Lares et al. 2001). One more plausible explanation for the inability of the sea urchin to depurate cadmium may be attributed to the incomplete recovery of the organism from the toxicity of the spiking exposure. If the metal is bound to cellular constituents it can potentially cause the dysfunction of enzymes, ion channels and DNA (Vijver et al. 2004). In this case the organism will be incapable of depurating cadmium no matter how clean the surrounding environment is. However, if toxic effects were occurring, I would not have expected the observed depuration of other contaminants such as lead, unless different detoxification pathways were in use to remove other toxicants.
The behaviour of nickel in this study is in agreement with the literature. It has been reported that nickel is not accumulated in the tissues in significant amounts by aquatic organisms (Gillis et al. 2004, ATSDR 2005) and there appears to be no evidence that nickel biomagnifies in aquatic food webs and, in fact, there is evidence to indicate that the nickel concentrations in organisms decrease with increasing trophic level (McGeer et al. 2003). It could be that nickel in its ionic form is not bioavailable to the roe of P. lividus as much as it is for the calcites in the test or it might be that nickel, when accumulated in tissues, is preferentially taken up in the gut via diet (Punt et al. 1998, Gillis et al. 2004, Hedouin et al. 2007). To test the bioavailability of nickel to calcites and its depuration, I examined the nickel in the test of P. lividus for both laboratory and field translocation depuration experiments using a semi quantitative method on ICP-MS (one metal nickel and one IS scandium). From this I observed a major increase in nickel concentration in the test after the exposure period (from 10 to 50 ppm), then fast depuration on the third day in both the laboratory and the field (down to 20 ppm). Depuration studies in the literature show that mussels and oysters can lose, up to 58% and 38% of nickel from their tissues respectively after 2 weeks in flowing seawater (ATSDR 2005). Even though the nickel depuration rate was not significant from day to day in our experiments, nevertheless it has been calculated for future reference.
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The reason behind the control urchins increasing their accumulation of vanadium to a comparable level to the concentration found in the toxicant exposed urchins is not very clear but it could potentially be due to vanadium migrating from other body parts (test and teeth perhaps). It is interesting to see that the final value reached by both these experiments on day 14 is below the baseline despite vanadium scoring the longest BHL time among the metals of the field depuration. It has perhaps to do with the way vanadium is sequestered in the tissue as discussed in the bioaccumulation chapter where it was seen that vanadium has shown weak penetration in the internal organs of benthic larvae (Fichet et al. 1998) and a depuration after 50 days. Amiard et al. ( 2004) also noted slow accumulation of vanadium in mussels which accumulated the vanadium after one month of the oil spill occuring. The slow accumulation was also mentioned by Venkataraman and Sudha (2005) who observed a poor absorption of vanadium from the gastrointestinal tract in humans and by Cebrian et al. (2003) who also noted an increase of vanadium in sponges in contaminated site but not enough to be significantly different.
The organic phase The depuration rate of dieldrin and 4,4‘DDE in the field translocation was characterised by a fast depuration phase followed by a slow phase. The initial rapid depuration phase of the OCPs observed in P. lividus could be due to the evacuation of unassimilated toxicants whereas the slower depuration phase represents the loss of toxicants that had been assimilated and incorporated into the tissues (Choi et al. 2003).
The sudden increase observed after the initial decrease observed in the laboratory depuration experiment (Figures 7 &8) may be due to the mobilisation of dieldrin and 4.4DDT (metabolised into DDE as seen in chapter 3) (ATSDR 2002b, Kwong et al. 2008) into the roe from the coelomic fluid. The coelomic fluid has the potential to accumulate organic and inorganic compounds and may act as a storage site for these toxicants. This increase in the OCPs concentrations during the depuration phase has been noted in many studies such as those conducted by Choi et al. (2003) and Antunes et al. (2007b). Those authors found that scallops (Chlamys nobilis), mussels (Perna viridis) and mullet (Mugil cephalus), depurate OCP after a 10-15 day starvation period and have elevated organochlorines levels in tissues during the depuration period. In this study, it is believed that a mobilization from coelomic fluid or digestive tract to the roe has occurred. In the laboratory depuration experiment I expect this exchange between coelomic fluid and the surrounding water medium to be slower than in the field translocation and that might explain why only the laboratory experiment showed this pattern and not the field experiments.
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Lipid consumption has been shown to be a factor driving the elimination kinetics of DDTs in the muscle tissue of mullet (Antunes et al. 2007b). Our sea urchins were fed only once a week and lipid consumption, due to the sea urchins starving six days a week, may have been a booster with regards to depuration. However, I would expect lipid consumption to result in lower weights but this was not the case in this study either because two weeks were not long enough to lose all lipids or because sea urchins were fed at least once a week.
Depuration rates of dieldrin and 4,4‘ DDE were comparable with existing studies on other marine organisms. 44% of total DDTs were eliminated from the whole body of marine fish after two weeks of depuration (Kwong et al. 2008) , 80% of DDT was eliminated from mullet M. cephalus (Antunes et al. 2007b), 54% lost in Gobius. Nudiceps (fish) within 3 days of exposure in fresh sea water (Wandiga et al. 2003). Depuration of 4,4‘DDE in P. lividus after two weeks was ca 66% in our field translocation study. As for dieldrin, 75% loss has been reported in the literature for green lipped mussels (Richardson et al. 2005) which is similar to the depuration rate seen in this study of ~ 66%.
In this case, the sea urchin is a good candidate for biomonitoring, not only because of its ability to bioaccumulate certain contaminants, but also with its ability to depurate, in case depuration kinetic studies are the major reason for this biomonitoring. However, I do not expect that the urchin would be a useful indicator of environmental loads of all contaminants. If P. lividus encounters a sudden exposure of lead (sudden pulse), sea urchin roe might accumulate lead but will depurate it as soon as the background sea water level of lead returns to ambient concentrations. If not sampled immediately, this information might be missed from the roe. Instead it is best to look for toxicants in other body parts like the coelomic fluid or the calcites because they have shown to store toxicants a bit longer than the roe especially outside spawning season. For cadmium, P.lividus, can be used any time during a scheduled yearly monitoring since the metal does not appear to depurate. For nickel, I recommend that the spines or test be analysed, instead of the roe, as the test and spines seem to be the preferred sites for nickel absorption. However, one must be careful when assessing the spines of sea urchins as these can be shed in times of stress, of spawning or if nitrates levels are high. Eventually, if the organism survives, new spines will develop (personal observation). The same applies to vanadium which even though it accumulates in calcites, its presence in roe is very low. Since dieldrin and 4,4‘DDE in the roe are removed quickly, coelomic fluid might be a better indicator of exposure to these contaminants.
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This study will hopefully be useful to the aquaculture industry as it will permit them to use the BHL as a general guideline for the maximum amount of time required for P. lividus to undergo depuration before being sold in the sea food market, at least for the toxicants I have tested. However, it is recommended that depuration studies be performed for each individual toxicant and not on classes of toxicants. This is because in some studies such as Gardinali et al. (2004) it was observed that depuration of PCB congeners in bivalves indicate that large differences in accumulation and depuration rates exist even among congeners of similar chlorination levels. This study has also shown that natural depuration (the cheapest and the more feasible), in more than one aspect, is much better than laboratory depuration and this will help the aquaculture industry cut their overhead costs.
Conclusion
Lead and vanadium both depurated from the roe whereas cadmium did not. This represents a problem for the future of sea urchin fisheries if cadmium levels were to exceed the safe levels set in food safety guidelines, as was the case in this study after exposure in the laboratory. EDTA was not a big booster for metal elimination because of the low concentration used in the study. Perhaps the EDTA should have been added in higher concentrations. Nickel was neither taken up in the roe in the two weeks exposure nor was it depurated back to its baseline concentration. 4,4‘DDE and dieldrin also depurated from the roe during the field translocation. The fluctuations seen in the laboratory depuration study might stem from metal exchange between the roe and coelomic fluid, unlike the coelomic fluid in field translocation, which appeared to depurate faster perhaps because of clean flowing water. In the case of metals, this fluctuation in the laboratory experiments could come from mobilization from the coelomic fluid, digestive tract or calcites. The sharp decline seen for some toxicants in the first days followed by a slower discharge to steady state is probably due to the elimination of toxicants not tightly bound and stemming from the sudden exposure of sea urchins to toxicants. Whereas the second slow phase of depuration may have been due to elimination of toxicants tightly bound at the cellular level.
This study has shown once more that the sea urchins are good bioaccumulators of certain contaminants. Since P. lividus only depurate when they are in a clean environment, its different body parts can be used in long and short-term biomonitoring studies. However, it is also important to take into account the characteristics of the toxicant such as its steady state and its common binding sites when performing these studies. Since
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depuration is toxicant specific (vs being classes of toxicants specific), this experiment must be repeated for all toxicants that are detected in the edible part of the P. lividus (mercury and copper to mention few), or in any other part that may be used in the production of pharmaceuticals. The use of radio labeled tracers is recommended since it helps highlight binding sites and will help to understand the mechanism of depuration. It is also recommended that some exposure to 4,4‘DDE and cadmium be undertaken to check on the possible additive effects these two toxicants might have on the hardness of the sea urchin skeleton. The BHL determined in this study will help the aquaculture industry meet food regulations pertaining to the sea urchin fishery.
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Chapter 5
5.Bioaccumulation of metals and PAHs in Australian sea urchin Heliocidaris erythrogramma
Abstract
The sea urchin Heliocidaris erythrogramma currently is part of a small wild harvest fishery in southeastern Australia. This study investigated the bioaccumulation in H. erythrogramma of contaminants associated with an oil spill. Metal levels were determined in gametes pre- and post- spawning. Differences in accumulation in male and female sea urchins were compared as well as background levels in H. erythrogramma collected from a site in the northern side of Sydney Harbour. Contaminant levels in the urchin roe (the edible part) were compared against international and national food safety regulations.
The metal baseline levels in the roe of the urchins collected from Sydney Harbour area were within the regulations except for lead which exceeded two international standards. Males and females differed in their bioaccumulation of metals. Whereas nickel and cadmium accumulated more in the former, lead was favoured by the latter and they both accumulated vanadium to the same level. Eggs shed during spawning carried high levels of metals, which may lead to reduced success in reproduction. Although the lipid content of females was double that of males, the adjusted values of PAHs (reported in lipid) in males suggest that males has accumulated more fluoranthene and pyrene than females. Only in acenaphthene did the female bioaccumulate more than the male.
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Chapter 5.Australia
Introduction
Many fisheries around the world are on the brink of collapse due to overharvesting of sea food products and the lack of sound management (Jacquet 2007, MacKenzie et al. 2009). At a local level, the loss of a traditional fishery can result in a major economic crisis (Roughgarden et al. 1996). Establishing non-traditional fisheries may therefore elevate pressure on already overharvested species while providing economic and social benefits to local communities. Heliocidaris erythrogramma, also known as the purple or green sea urchin (Valenciennes 1846), is an Australian endemic echinoderm (Pollard et al. 2002). It can be found on rocky reefs from the intertidal zone down to approx. 35 m along the east, west and south Australian coasts (Kailola et al. 1993) but is mainly present in shallow waters (Underwood 1991). As benthic animals, they are often found attached to rocky reefs, stones, seagrass beds, in crevices and burrows. The species is common throughout its distribution and in some areas may reach densities of ca. 20-80 individuals per square meter (Andrew 1999). Wright and Steinberg (2001) have also observed higher densities of 80-192 individuals per square meter in 3-4 m depth in a subtidal macroalgal community near Sydney NSW and at such densities can cause shifts from foliose to crustose-algal dominated ecosystems (Wright et al. 2005). While usually 60-90 mm in diameter H. erythrogramma can reach 125 mm in Tasmania (Growns et al. 1994). Despite large numbers of H. erythrogramma in Australia, they have so far supported only small wild harvest fisheries in various states within the country but these have the potential to expand. Annual catches of H. erythrogramma in Victorian waters ranged from 19 to 50 tonnes prior to 2004, and has been fairly constant at ~20 tonnes since (DPI Victoria, pers. comm.). The Tasmanian H. erythrogramma fishery is the largest in Australia, landing over 42 tonnes in the 2009/10 quota year (Fisheries 2010).In New South Wales, there are restricted landings (less than 500kg per year), however the congener Heliocidaris tuberculata has been preferentially targeted for extraction due to its better quality roe (Heasman et al. 2007).
In order to expand and enhance the H. erythrogramma fishery in Australia, there is a need to understand, not only the abundance, life history and population demography of the species, but also the bioaccumulation and biomagnification processes that occur in this species and its compliance to food safety regulations. Testing the bioaccumulation of petroleum-based compounds is particularly pertinent in the marine environment due to the ubiquitous nature of these compounds. Here, oil releases are a result of human activity and natural seepage out of fissures and cracks in the sea floor, especially after earth tremors (Suchanek 1993, UNEP et al. 1995). Regardless of the source of the oil, or even the nature of the release (i.e. sudden spill or chronic release), oil is generally perceived as a major threat to the environment that
123 Chapter 5.Australia can possibly affect the viability of a species as an expanding fishery (UNEP et al. 1995). The toxicological effects of oil on marine organisms have the potential to cause mortality and deformity (Hartung 1995, Michel et al. 1997), while toxic components of the oil (e.g. Polycyclic Aromatic Hydrocarbons and metals) may be accumulated (Brandt et al. 2002, Kelly et al. 2008) they do not biomagnify through the food web ( Gray 2002) . Reproduction (formation of gametes, fertilization and hatching success) and growth may also be impaired due to the toxic effects of oil at the cellular level, which may then affect population and community processes (Sol et al. 2000, Pillai et al. 2003, Velando et al. 2005, Bellas et al. 2007). H. erythrogramma has been subject of many ecological and biological studies (Pederson et al. 2008) (Laegdsgaard et al. 1991, Huggett et al. 2006), but to the best of my knowledge, no bioaccumulation studies have ever been done on H. erythrogramma.
There are a number of factors known to influence bioconcentration in a suite of marine invertebrates, such as age, size, seasonality, growth rates, genetic and phenotypic variation (Burger et al. 2007). One of the most interesting and widely debated causes of differences in bioaccumulation is sex-dependent accumulation in tissues (Couceiro et al. 2009). Part of this interest lies in whether gametes (eggs in female, sperm in males) will accumulate metals differently (Akberali et al. 1985) and whether this accumulation will impact the fertility or ongoing health of the next generation. In fact, works by Russian researchers in the late 80s have shown that a 50 ug/L of cadmium in water for 15 days, has led to the production of abnormal gametes and non-viable off spring in Strongylocentrotus intermedius (Khristoforova et al. 1984, Gnezdilova et al. 1987). Others researchers have noted that even if the fertilization to birth process was successful, defect can still be seen later in the progeny (Fichet et al. 1998) which might be born with denatured DNA.
Gender differences in bioaccumulation may also be mediated by the content of lipids. Lipids are a group of organic compounds which include fats, oils, waxes, sterols, and triglycerides (Van Geest et al. 2010). In some species, lipid content has been observed to be higher in females (McBride et al. 1998). This may lead to differences, not only in the accumulation of PAH type compounds in the H. erythragramma fat store, but can also lead to spurious conclusions made from measurements of total body bioaccumulation measures. If females have more body fat than males then the organic toxicants burden in the organism must be expressed in lipids rather in total weight, if gender comparisons are to be made (Burger et al. 2007).
This study represents an initial project investigating bioaccumulation of compounds such as be petro chemical and heavy metal contaminants in H. erythrogramma. I first determined metal levels in the roe pre- and post- spawning. The study reports on differences in 124 Chapter 5.Australia accumulation in male and female sea urchins and reports background levels in H.erythrogramma collected from a site in outer Sydney Harbour. Contaminant levels in the urchin roe (edible) are compared against international and national food safety regulations.
Methods
To measure the response of the sea urchin Heliocidaris erythrogramma to a probable scenario of an oil spill accident, an experiment was undertaken in which Heliocidaris erythrogramma (Valenciennes, 1846) were exposed to three PAHs (pyrene, Fluoranthene and acenaphthene) and four metals (nickel, vanadium, lead and cadmium) in ratios estimated to be equivalent to that following the release of crude oil into the sea.
Urchin Collection H. erythrogramma urchins were collected from shallow subtidal rocky reef at Fairlight Beach, Sydney, Australia. Fairlight Beach (33° 47′ 45.6″ S, 151° 16′ 37.2″ E-33.796, 151.277) is located on the Northern shore of Sydney Harbour (Port Jackson), relatively close to the estuary mouth. Land uses within the Manly beach catchments include residential, commercial, industrial, recreational and bushland (SHOROC 2010). Urban run off reaches the sea via storm water drains and sewage outfalls. This area is known to have high metal levels in the sediments, exceeding the minimum effect range on biota (Birch et al. 2002) as highlighted by ANZECC Sediment Quality Guidelines-Low and -High (ISQG-L and –H) ranges (ANZECC 2000a). Urchins were collected from forests of the alga Ecklonia radiata by scuba diving off Fairlight (2 to 5 m depth) on the 16th of November 2006 (Figure 1).
Exposure
Two treatments were established (control and contaminated) in eight aquaria (four replicate aquaria for each treatment). Each contaminated treatment tanks consisted of 40 L filtered seawater spiked with a 50 ml solution of deionized water:acetone 50:50 and containing 1 µg g-1 of each of fluoranthene, pyrene, acenaphthene and 0.1 µg g-1 of nickel, vanadium, cadmium and lead. The control tanks consisted of 40 L of filtered seawater spiked with 50 ml of a 50:50 mixture of Deionized water:acetone free of toxicants. The choice for the first two metals was based on the observed toxic effects on reproduction of sea urchins and because they are bi-product of industrial activities (Au et al. 2003, Xu et al. 2010). Vanadium and nickel were chosen because these two metals are usually found in the dispersed fuel oil (Tronczynski et al. 2004) and are regulated for their maximum presence in petroleum derivatives . PAHs (and other aromatic petroleum hydrocarbons) are components and markers of crude oil and they are pollutants of concern because of their toxic and carcinogenic potential (ATSDR 1995) when accumulated in biota.
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Spiking preparations The three PAHs (fluoranthene, pyrene and acenaphthene,high purity) were purchased from Fluka and Merck and diluted in their proper dissolving solvents before being added to the carrier solvent (acetone). Inorganic metal toxicants (vanadium, nickel, cadmium and lead), were also purchased from Agilent, were diluted in deionized water (acidified with HCl) before being added to the acetone fraction. As discussed above, this cocktail of toxicants was an attempt to simulate as close as possible the real environmental conditions of an oil spill.
126 Chapter 5.Australia
Figure 1. a- map of Sydney- New South Wales-Australia (New South Wales Highlighted), b- map of Sydney Harbour and c-map of Fairlight-Sydney-Australia.
127 Chapter 5.Australia
Experiment 20 urchins were placed in each aquarium. The size of sea urchin ranged from 4 to 5 cm (measured using a Vernier caliper). The sea urchins were fed ad libitum the algae Ecklonia radiata and were left to acclimatize for four days before the start of experiment. As soon as the sea urchins were received in the laboratory, 10 sea urchins (mixed gender) were randomly chosen from the tanks, excised into different body parts and processed for metal analysis. These are referred to as the baseline.
The experiment lasted for three weeks and four days: from the 16th of November till 11th of December. Water was changed every five days. Fresh filtered water was discharged into the tanks, spiking of the water occurred straight after the change of water. In the first week of December, two weeks into the experiment one control tank and one treatment tank were lost due to a sudden unexplained deterioration in water quality.
In summary, the experiment had three consecutive steps: acclimatization (also considered as depuration period for sea urchins), exposure and spawning. Acclimatisation lasted four days and was also considered as a depuration time for the sea urchins (mixed gender) to purge from their roe any toxicants accumulated in the field (the metal baseline in all body parts of sea urchins has already been established in the laboratory as mentioned previously at day 0). For the sea urchins in control tanks, 4 days of acclimatisation and 21 days in clean water can be considered as depuration period. The second step was the exposure to toxicant which spanned 21 days. At the end of this period, 3 urchins (mixed gender) were collected from each tank to be tested for toxicants bioaccumulation in the roe of both control and spiked treatments. The third step was the spawning done on the remaining of the sea urchins in the tanks and it lasted one day. At the end of it, 3 pairs of males and females from each tank were selected for the analysis of contamination in the roe and the eggs.
At the end of the three week exposure period, three sea urchins were taken out of each tank (n=3) to analyse for bioaccumulation (they are referred to as sea urchins after 21 day before spawning), the fifteen remaining sea urchins were taken out of the tanks and put on specialised tray to undergo spawning. These were induced to release their eggs or sperm by injecting 2ml of 1M KCl into their coelomic cavity. Once the 3 males and 3 females from each tank were selected, they were dissected before undergoing further gender confirmation under microscope. Their roe was then extracted and tested for differences in bioaccumulation between genders (see Analytical techniques)(they are referred to as male/female after 21 days after spawning). The eggs released by the female sea urchins were collected on a mesh, washed with 1% nitric acid, freeze dried, digested in a microwave and analyzed for metal content in order to assess if purging of toxicants has occurred during spawning (transfer of metal from parents to offspring) (see Analytical techniques).
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Analytical Techniques
Metals The metal digestion was done at the Analytical Laboratories, Chemistry department, University of New South Wales. Roe samples and eggs were freeze-dried. Spines, tests and teeth were washed with 1% nitric acid then oven dried. The roe and eggs were weighed and digested in a closed microwave using nitric acid and peroxide for roe, and nitric acid for the rest of the body parts. The quality control protocol, used in the digestion, consisted of the analysis of a blank, a standard reference material (NIST oyster tissue 1566b, bovine 1577), duplicates and matrix spikes. The protocol is fully described in the temporal variation experiment.
To check for sample preparation and analysis errors and as a mean of comparisons, the roe samples digested in Australia were split into two sets and all the samples of the second set were transferred to Lebanon and were analysed at the ISO 17025 accredited Environment Core Laboratory at the American University of Beirut. Digests were run on the ion coupled plasma-mass spectrometer (ICP-MS 7500ce-Agilent equipped with a Cell Dynamic Range and an autosampler) for cadmium, lead, nickel and vanadium determination. The quality control protocol used when running samples on ICP was done as described in the temporal variation chapter. An internal standard mixture consisting of scandium, thallium, ytterbium and cerium was added to all samples and all calibration levels for the set done in Lebanon. Rhodium was used in addition to the above mentioned IS for the set run in Australia. Recoveries for all quality control samples were accepted when values were within the control limit charts of the corresponding metals (80-120% on average). The generated data was treated for both the digestion and analyses and calculated from both external standard curve and ratio of Internal Standard (IS) using Chemstation (B.03.07). Data for lead in the roe (25% of the data) was further treated for recovery when the CRM was not within the limit. Results were expressed as microgram of metal per g of dry weight (µg g-1). The quantification limit of the tissues for all metals is 0.01 µg g-1. For statistical purposes values less than the detection limit were reported as half the limit of quantification. All methods used for digestion and analysis were from international approved methods with some modifications (EPA-USA-200.8 1994, EPA-USA-3052 1996, APHA 1999).
PAH and Lipids
Microwave Extraction
Sample extractions were performed using an Ethos 1 (1000 W) Microwave Extraction System (Milestone) equipped with a stirrer, temperature sensor and 10 vessels with a nominal volume of 100 mL. Temperature and pressure inside the extraction vessels can
129 Chapter 5.Australia reach a maximum of 260 C and 35bar (500 psi), respectively. The glassware was washed with detergent, rinsed with DI water and acetone (laboratory-grade) before being heated in an oven at 180 oC for 2h. A 0.1 g of homogenised roe was weighed and put in a glass extraction vessel designed for organic extraction. A surrogate standard Phenanthrene d-10 (1 mg l-1 certified) was added to the roe directly, mixed and left to settle for half an hour. 3 ml of methanolic potassium hydroxide, and 10ml hexane-acetone (1:1) were added to the roe in the vessel before closing it for digestion. In preparation for column clean up, the final solvent extract was filtered and concentrated to 1ml using the Ethos post-extraction solvent evaporation apparatus equipped with a vaccum pump. All solvents used were of chromatography grade and purchased from Merck, Acros Organics, Riedel-DeHaen, Fisher and Fluka. During the analysis, quality control protocol (QC) was carried out on each set of samples to check performance of digestion (blank of digestion, spiked blanks, duplicates, sample spiked).
Lipid determination was done according to a method provided by manufacturer of the microwave digestor (Ethos-Milestone) and stored in its system. It consists briefly of a microwave digestion where 1 g of roe was weighed then extracted with petroleum ether:acetone (80:20) for 20 min at 1000 W and 90oC. The organic sample extract was then filtered before being evaporated to dryness and weighed again for lipid determination, which is expressed in % of weight.
Clean up
As a clean-up step, silica, florisil and sodium sulphate (2:1:2) were used according to the protocol described in the bioaccumulation chapter. Samples were later analysed on on gas chromatography- mass spectrometry ( GC-MS) scan and selective ion monitor (SIM) modes for PAH analysis and confirmation.
Quality Control
All instruments used in this study have undergone an Operational Qualification and Verification of Performance (OQVP). On the day of use, the chromatography instruments were tuned, maintained and checked for good performance using certified quality standards purchased from Agilent. All internal and external standards used for calculating response factor for PAHs have a purity of more than 99% and were purchased from Agilent along with certificates. Results of roe are not expressed in µg of organic toxicant g-1 of lipid but rather in µg of organic toxicant g-1 of dry weight of roe. The quantification limit of the tissues for PAHs is 0.1 µg g-1.
Peak identification and confirmation for both organic categories were performed on the samples using a HP 6890 Series GC-MS 5975c series equipped with an AT-5MS (30 m 130 Chapter 5.Australia length, 0.25 mm diameter and 0.25 µm thickness) column, G1513A autosampler ,operated at the electron impact mode (EIM) and using Chemstation software (D.03.00) for data acquisition and quantification. Mass spectra were acquired using both the scan and the SIM modes. In the scan mode, and due to the complicated chromatographic features associated with the roe complex, matrix components were identified by comparison of their mass spectra with the Wiley mass spectra library (version 1.4) and their retention time. In the SIM mode, three fragmentation ions (When possible as some compounds have less than two) for each toxicant were chosen for mass scanning. The dwell time was 50 ms for windows. Temperature program was as follows: from 40 °C to 160 °C at a rate of 5°C /min then to 250°C at a rate of 3°C /min and finally to 300°C at a rate of 25°C /min and constant for 10 min (total run time was 70min). Split-splitless mode inlet was at 250 °C Helium gas was run with a total flow of 18 ml/min, MSD transfer line at 300 °C, MS Quad 150°C and MS source 230 °C.
Statistical Analysis Differences in the size (diameter in cm), roe weight (g of wet weight) and total weight (g of wet weight) of the mixed gender for the two treatments (Control and Contaminated) were analysed to eliminate these parameters as cofactors in the bioaccumulation experiment using permutation analysis (PERMANOVA). Lipid differences between males and females, in each of the treatments, were tested using One Factor ANOVA. Lipid differences, of mixed genders between control and treatment, were tested using One Factor ANOVA.
Differences in metal content of eggs between treatment and control were assessed using a single factor ANOVA. One Factor ANOVA was used to check for differences in bioaccumulation between different body parts of the sea urchins collected from the field (baseline). A One Factor ANOVA was also used to test difference in bioaccumulation in the roe between control and treatment. It was also used to check differences for each gender between control and treatment.
Because males and females were sampled from the same tank, I did no formal analysis of gender differences and our observations are based on graphical presentation only. To test for differences in bioaccumulation of PAHs between control and contaminated treatments in males and in females separately, a non-parametric Kruskal-Walis analysis was used when ANOVA assumptions could not be met. I also conducted a One Factor analysis, using Permutational analysis (PERMANOVA), for which no assumptions were satisfied, to confirm results. For the sake of brevity only the non-parametric results are presented.
Whenever ANOVA was used, assumptions of Normality (tested with P-P plot) and Homogeneity of variances (using Levene‘s test) were satisfied before analysis. SPSS(2010)
131 Chapter 5.Australia software (version 18), compatible with IBM and Primer (v6) (Clarke et al. 2006) were used for all statistical analyses.
Results
Size , weights, lipid percentage The size (diameter of test without spines), height (oral-aboral), total wet weight and the roe wet weight of the sea urchins did not differ between treatments (control and contaminated) or between gender for the three weeks of experiments (p>0.05 ). Therefore none of these parameters were considered as cofactors in the statistical analyses (Figure 2).
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Figure 2. a-Size (test diameter) and height (oral-aboral) in cm (±SE) of male and female Heliocidaris erythrogramma used in the bioaccumulation experiment. b- Size (diameter in cm±SE) of mixed gender of H. Erythrogramma Baseline data on arrival, data after the three weeks of exposure to two treatments control and contamination, c-Roe wet weight of sea urchin in g (±SE) of mixed gender of H. Erythrogramma Baseline data on arrival, data after the three weeks of exposure to two treatments control and contamination, d-Total wet weight of sea urchin in g (±SE) of mixed gender of H. erythrogramma on arrival and after three weeks of exposure to two treatments control and contamination.
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Figure 3. a-Total lipid percentage (±SE) in roe of mixed gender of Heliocidaris erythrogramma used in the bioaccumulation experiment gender on arrival to laboratory (baseline) and after the three weeks experiments for both treatments control and contamination, before spawning, b- Total lipid percentage (±SE) in roe of males and females of H. erythrogramma used in the bioaccumulation experiment at the end of the experiment for both treatments control and contaminant and after spawning.
No differences in size were observed between the mixed gender (sexes mixed and undifferentiated) and separate gender (male differentiated from females) (figure 2 a and b), ranging from 4.3 to 5 cm. thein mixed gender treatments, the total weight was around 40 to 45 g while the roe weight increased from 9g to 12g after 3 weeks of toxicant exposure in one of the treatments and after feeding ad libitum (Figure 2).
The roe lipid content of females was almost double the amount of the lipid measured in males (p=0.011)(Figure 3 a). In mixed gender, the lipid concentrations in control and treatment did not differ (p>0.05) and both were lower than the concentrations of separate gender (Figure 3b). Levels of fluoranthene were adjusted in lipid of both gender and concentrations became 595 µg g-1 in male and 270 µg g-1 in females. As for pyrene, after the lipid adjustment, the concentrations in males and females were 430 and 192 µg g-1 respectively.
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Metals in urchins upon collection All baseline metal concentrations except lead, were greatest in the roe followed by the test, spines and teeth. Lead was the highest in the test, teeth then roe and spines.
Figure 4. Concentrations (±SE) in µg g-1 of nickel, cadmium, vanadiuml and lead in the different body parts (roe, test, spines,teeth dryweight) of Heliocidaris erythrogramma directly after collection from the field. The significance of One Factor ANOVA at 0.05 is enclosed in addition to the Tukey‘s post hoc test. Dotted lines over body parts mean no significant difference between those treatments.
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Metals in roe and eggs Following exposure to contaminants, H. erythrogramma accumulated metals in the roe to levels much greater than baseline or control treatments (fig 5). In urchins from the contaminant treatment, the concentration of vanadium in the roe was almost triple the concentration in the control (mixed gender, table 1, figure 5). No gender differences in Vanadium concentrations were observed (Figure5a).
Figure 5. Concentration in µg g-1 (±SE) of vanadium, nickel, cadmium and lead in the roe (dry weight) of H. erythrogramma for mixed gender on arrival (baseline) and after 21 days of exposure to both treatments (control and spiked) , also concentration in male and female of the sea urchin after 21 days of exposure for both treatments (control and spiked) but after spawning.
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Table 1. ANOVA results of the exposure experiment of nickel, cadmium, lead and vanadium (µg of metal per g of dry weight)(µg g-1±SE) in roe of Heliocidaris erythrogramma.(significance at 0.05), analysing the mixed gender before (baseline) and after 21 days of spiking for both control and spiked treatments. Also comparing for each gender (male and female), the difference between the treatments (control vs spiked).
One Factor
Mixed gender Vanadium Cadmium Nickel Lead Transformation Nil Nil Nil Nil
Distribution Normal Normal Log Normal
Levene 0.405 0.087 0.06 0.3
df 1 1 1 1
MS 13.2 7.992 1.463 0.753
p (Control vs Spike treatment) 0.001 0.008 0.044 0.015 Male Transformation Nil Nil Nil Nil Distribution Normal Normal Normal Normal
Levene 0.051 0.123 0.393 0.083
df 1 1 1 1
MS 20.983 14.04 6.337 0.063
p (Control vs Spike treatment) 0.175 0.000 0.03 0.4 Female Transformation Nil Nil Nil Nil Distribution Normal Normal Normal Normal
Levene 0.122 0.591 0.307 0.552
df 1 1 1 1
MS 23.72 4.649 0.597 0.101
p (Control vs Spike treatment) 0.17 0.001 0.192 0.672
Cadmium accumulation in males was 33% higher than in females. In the mixed gender, concentrations increased from 0.2 ug.g-1 in control urchins to 3 ug.g-1 in urchins from the spiked treatment.(Figure 5b). Nickel accumulation in male was 35% higher than in females with treatment in mixed gender reaching after 3 weeks 2.5 times the baseline (Figure 5c). The lead accumulation in the roe did not differ much between control and treatment at the separate gender level, between control and spiked treatments, however females accumulation was around 33% more than males (Figure 5d). In the mixed gender, a small increase from the control was noticed in the spiked treatment.
Vanadium concentration in eggs of Heliocidaris erythrogramma was 5 fold higher in the contaminant exposure treatment than in the control (Figure 6a). The concentration in the eggs was only 10% of the total concentration of V in the roe. Cadmium concentration in the eggs was equivalent to that in the roe, however was 20 times more in the toxicant exposure treatment than in control (Figure 6b). Nickel concentrations were 2.5 times higher in urchins from the contaminant treatments than control (Figure6 c). The lead concentration of eggs in spiked treatment was double that of control and double that of the total roe (Figure 6d). 137 Chapter 5.Australia
Figure 6. A Contaminant concentrations (Mean±SE) in µg g-1 of vanadium, nickel, cadmium and lead in the eggs (dry weight) of Heliocidaris erythrogramma after 21 days of exposure to both treatments (control and spiked). Results of One Factor ANOVA (α-0.05, control vs spiked treatment) are enclosed in the figure.
PAHs in roe PAH accumulation in the roe of H. erythrogramma differed depending on the analyte measured. The control tissue concentrations were below 0.1 ug.g-1 for all PAHs. The highest accumulation occurred for fluoranthene (Figure 7a) where exposed urchin roe reached 140 ug.g-1 while pyrene concentrations reached 100 ug.g-1 (Figure 7c) . Females accumulated 12 and 15% more than males of pyrene and fluoranthene respectively. For acenaphthene, the accumulation was less pronounced, reaching 5 ug.g-1in females, while the males did not accumulate above 0.1 ug.g-1(Figure 7b).
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Figure 7. Concentration of fluoranthene, acenaphthene and pyrene in µg g-1 (±SE) in the roe (dry weight) of Heliocidaris erythrogramma for mixed gender on arrival (baseline) and after 21 days of exposure to both treatments (control and spiked) , also concentration in male and female of the sea urchin after 21 days of exposure to both treatments (control and spiked) but after spawning.
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Table 2. ANOVA results of the exposure experiment to fluoranthene, pyrene and acenaphthene (µg of PAH per g of dry weight)(µg g-1±SE) in roe of Heliocidaris erythrogramma.(significance at 0.05), analysing the mixed gender before and after toxicants being added (control vs spiked treatments) then comparing for each of the two genders (male and female) the difference between the treatments (control vs spiked treatment). When assumptions of ANOVA were not met, a non-parametric test (Kruskal –Wallis test) was used to detect differences between treatments and the p value would be annoted as such.
Mixed gender Fluoranthene Pyrene Acenaphthene Transformation Nil Not tested Not tested
Distribution Normal Not tested Not tested
Levene 0.055 Not tested Not tested
df 1 1 1
MS 27208.15 ------0.037 non 0.037non p (Control vs Spike treatment) 0.000 parametric parametric Male Transformation Not tested Nil Not tested Distribution Not tested Normal Not tested
Levene Not tested 0.113 Not tested
df 1 1 1
MS --- 11162.9 --- 0.037 non 0.025 non p (Control vs Spike treatment) parametric 0.042 parametric Female Transformation Nil Not tested Nil Distribution Normal Not tested Normal
Levene 0.064 Not tested 0.08
df 1 1 1
MS 27431.9 --- 85.779 0.037 non p (Control vs Spike treatment) 0.005 parametric 0.035
Discussion
In the field, the sea urchin Heliocidaris erythrogramma had measurable levels of nickel, cadmium, lead and vanadium in the roe, test, spines and teeth. Urchins have rapidly bioaccumulated metals and organics in the roe when they were exposed to a representative cocktail of contaminants in the laboratory. Cadmium and nickel were accumulated in higher levels in males, while no gender differences in vanadium concentrations were observed. Female urchins accumulated more lead than males. After lipid adjustment, males accumulated more PAHs than females, with concentrations far above ambient levels.
When setting up biomonitoring programs for new or emerging fisheries the focus should be largely on the edible part of any organism. In the case of sea urchins, the edible part is the roe. Metal levels in urchin roe recorded in this study are usefully compared to four international safety food regulations: the FDA-USA (2001) (Food and Drug Agency-Unites
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States of America), CODEX/FAO/WHO (2009) (Codex Alimentarius - Food Agency Organisation - World Health Organisation), The EU (2006) (European Union) and Australian food safety standard (ANSTAT 2011) (Table 1). Unfortunately the vanadium does not yet have any established food safety standards. These guidelines are more generally applied to fish and crustaceans; however they can, and are, commonly used as standards for comparison with other marine organisms. The baseline levels of nickel and cadmium were well under all the international regulations. However, concentrations of lead, were above the Codex Alimentarius/WHO and Australian regulations, and were close to being dangerously high according the FDA and EU standards. Lead is toxic to marine invertebrates and humans, with the potential to cause a range of symptoms such as physiological perturbations, growth suppression and reproductive damage in marine organisms and in humans (ATSDR 2007, 2008, Albanese et al. 2008 , García-Pérez et al. 2010). Elevated lead levels in the roe of H. erythrogramma should prompt further investigation especially in light of the fishery of this species. The effect of this lead level must also be considered when assessing reproduction in H. erythrogramma especially on the gametes and viability of off springs. Tests, similar to the test undertaken by Russian authors (Khristoforova et al. 1984, Gnezdilova et al. 1987) must be undertaken to see if this level can also affect the gametogenesis an the development of the organism from a larvae to an adult sea urchin in case fertilisation was successful. Lead will most probably cause physiological perturbations, growth suppression and reproductive damage in marine organisms and in humans (ATSDR 2007).
In order to determine the possible sources of high lead content in the roe of the sea urchin, background metal levels in both food and water would be required. There is a paucity of information on metal levels in the seawater of Sydney Harbour. The popular ‗Harbour Watch‘ program, currently undertaken on seawater within the Harbour, is mostly limited to the monitoring of fecal and total coliforms (DepartmentofEnvironment 2003). Some substantial sediment monitoring has been conducted previously by Birch and Taylor (2002). Here, Sydney harbour sediments were found to have a 35 to 67% probability of being toxic to biota (ANZECC guidelines). Birch recorded a range of 200 to 291 ug .g-1 of lead in sediments off Fairlight beach area. Birch and Taylor (2002) concluded that the sediments closest to the storm water canal had the highest metal concentrations. Similarly, studies conducted further inland, found the lead concentrations less than 0.4 µg l-1 in water, 242 µg g-1 in sediment and 61 to 264 µg g-1 in macroalgae (Melville et al. 2007).. It can be assumed that the macroalgae of Fairlight might have similar values as those of further up the estuary. I suggest that the high metal concentrations in the algae and sediments are a potential origin of the lead concentrations found in the sea urchins in this study.
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Table 3. Comparing the metal baseline concentration (mean µg of metal in g of dry weight) in roe of Heliocidaris erythrogramma of Fairlight to the different international food regulation standards. Besides each standard is the name of the organism for which the regulation was set. In case of different organisms presents, a range of these values are mentioned alongside a notice ―assorted marine organisms‖. No international food standard regulation has been set yet for Vanadium.
Metals Sea urchin CODEX/WHO/FAO Organism FDA- Organism EU standard Organism Australian Organism baseline level standard regulated USA regulated Mean µg g-1 regulated standard regulated Fairlight, Mean µg g-1 standard Mean µg NSW Mean µg g-1 Australia g-1 Mean µg g-1 assorted marine Lead 1.3 0.3 fish 1.5 Crustaceans 0.5-1.5 organisms 0.5 Fish assorted marine Cadmium 0.13 2 fish 3 Crustaceans 0.5-1 organisms 2 Molluscs
Nickel 1 N/A N/A 70 Crustaceans N/A N/A N/A N/A
Vanadium 2 N/A N/A N/A N/A N/A N/A N/A N/A
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Most metals, except lead, were more concentrated in the roe than in other body parts of the sea urchin. In general, increased metal uptake in urchin soft tissues has been observed in previous studies and experiments, especially in mature roe (Miramand et al. 1980, Warnau et al. 1997). Unpredictably I found decreased lead concentrations in the urchin roe in the current study. This may have been caused by the fact that urchins underwent spiking treatments when roe was immature or in regression (outside spawning season) but unlikely as the main spawning season for this species occurs from late November to March along eastern Australia (Williams et al. 1975, Laegdsgaard et al. 1991). Alternatively, biochemical properties of the roe may inhibit lead accumulation. This last observation is consistent with our observations from the laboratory exposure experiment, where lead concentrations in the urchin roe scored the lowest percentage bioaccumulation relative to other metals. Interestingly, the lead concentration was assessed during the pregametogenic stage of the gonads in the females (October-November for H. erythrogramma), when the roe was full of eggs and containing the maximum nutrients. This is the preferred stage for roe consumption in sea urchins (Lawrence 2007) and fetches the highest price at the markets.
I found that metal accumulation in H. Erythrogramma was gender related and metal specific. Nickel and cadmium were accumulated more in the male than in female urchins whereas lead was more heavily concentrated by females. Some studies have also observed gender differences in bioaccumulation. For example, cadmium and lead concentrations in female rock crabs are significantly higher than in males (Chen et al. 2005) and in sea star A. rubens (Temara et al. 2002). For copper and cadmium, body concentrations in females amphipods were greater than males (Marsden et al. 2003), while uranium in male and female fresh water bivalves did not differ (Markich 2003). This was also the case for vanadium in this study where I did not record any difference in bioaccumulation between males and females when looking at external features. In general, it is not possible to distinguish male from female H.eyrthrogramma by external examination. This is a potential issue since it is difficult to target sexes for accumulation monitoring should this become an issue in the fishery. Studies on sea urchin P. lividus have shown that the difference in accumulation between gender stems from the difference in their immune cells where females was found to have more phagocytes than males (Arizza et al. 2013).
When the metal baselines (i.e. day 0 at arrival, before the application of any treatment) of sea urchin mixed gender were compared to the sea urchin from the control tanks (after 4 days of depuration and 21 days of incubation in clean water) there was no depuration witnessed for vanadium or nickel. On the other hand, a slight increase was noticed in the cadmium and lead. This increase could have stemmed from the algae (Ecklonia radiata) fed ad libitum and which were collected from the same area where the sea urchins were fished. The algae may have accumulated cadmium and lead (Melville et al. 2007) which in turn could easily be biomagnified in the food chain. This does not rule out that algae might also have accumulated nickel and vanadium but their uptake into the roe is nevertheless 143 Chapter 5.Australia very slow (Chapter 4) and takes time to occur. The accumulated metals in the sea urchins of the baseline treatment, are a product of a chronic and low exposure to metals. Metals that are taken up in a sudden and acute exposure can be eliminated after 25 days of being in clean water as seen in depuration of sea urchins (chapter 4).
It is expected that both sexes would have a variety of methods of ridding the body of contaminants. Females, for instance, can rid their body of contaminants through shedding eggs (Temara et al. 1997, Burger et al. 2007). In terms of metal being accumulated in the eggs, it is quite obvious that metals in our study have been transmitted to the gametes in high quantities. From this study it would appear that there was a relatively slow process for lead and cadmium to reach the gametes whereas the accumulation of the vanadium and nickel in gametes was quicker. Metal transmission to gametes has the potential for significant consequences on urchin reproduction. For instance, the toxic effects of vanadium, lead and cadmium on larvae of another urchin, P lividus starts from their presence in sea water at concentrations equivalent to 100 µg l-1(Fichet et al. 1998, Radenac et al. 2001). Sometimes, the presence of toxic metals in biota may not hinder the development of urchins embryos, but it causes retards and malformations of the skeleton and gut of larvae as noted by Accornero and Manfra (2004) for the P. lividus species.
Not only was H. erythrogramma a rapid bioaccumulator of metals but it also accumulated significant amounts of PAH compounds. At first glance it may be thought that although the difference in lipid accumulation between male and females is very large, it did not affect the PAH accumulation between the genders. Since PAHs are lipophilic and are expected to accumulate in lipid, it was expected that females would accumulate more. However, if I adjust the concentration of PAHs to the lipid concentrations rather than in the total dry weight of the roe, it will show that in fact males have accumulated almost double the levels accumulated by females (levels of fluoranthene adjusted in lipid: in male 595 µg g-1, female 270 µg g-1; pyrene: male 430 and female 192 µg g-1) and that the female lipid seems to have effectively diluted the PAHs. It may also mean that the lipid concentrations in males did not reach a saturation concentration with regards to the accumulation of PAHs. Only acenaphthene was taken up by females of sea urchins H. Erythrogramma in levels much higher than in males after the adjustment relative to roe lipid content.
Antunes et al. (2007a) attributed the high levels of organics (PCB) in males of sardines to the slower mobility of these compounds in lipids as the lipids were consumed during the final period of spawning. Even with an equal lipid percentage between males and females, Hughes et al.(2006) showed that the difference between mature males and female sea urchins might be due to significant differences in the fatty acid composition of their gonads, as post spawned individuals showed no gender differences. Hughes et al (2006) also show that male urchins had higher levels of polyunsaturated fatty acids compared to females and that there was a dramatic reduction in the fatty
144 Chapter 5.Australia acids 22:6 (n-3) and 20:5(n-3) with increasing maturity stage. Investigating the correlation between fatty acid content and bioaccumulation may be a promising issue to investigate in the future.
The high levels of PAHs accumulated in the roe might have toxic effects on the urchins similar to the one reported by Schafer and Kohler (2009) who noticed gonadal lesions of female sea urchin Psammechinus miliaris after exposure to 500 µg l-1 of the PAH phenanthrene over 20 days. Similarly, Tilgham Hall and Oris (1991) noticed a decreased reproductive output (number of eggs laid) in fish exposed to 20 µg l-1 of anthracene for three weeks. One can then conclude that if an oil spill, similar to the one that occurred in Lebanon in 2006, (where concentrations of more than 1 mg l-1 of PAHs were reached in the sea) ever occurs in the Fairlight area, it would have devastating effects on the population of the sea urchins especially on reproduction.
Conclusion The sea urchin Heliocidaris erythrogramma is a potential bioindicator to be used in biomonitoring programs before its fisheries can be fully established. The metals‘ baseline levels in the roe of the urchins collected from Fairlight area were within the regulations except for lead which exceeded two of the standards and almost exceeded two other international regulations. Males and females differed in their bioaccumulation of metals. Whereas nickel and cadmium were accumulated more in the male than in female, accumulation of lead increased in females. Both males and females accumulated vanadium. Eggs were found to have accumulated high levels of metals, which may lead to reduced success in reproduction. Although the lipid in females was double that of males, the adjusted values of PAHs (reported in lipid) in males suggest that males had accumulated more fluoranthene and pyrene than females. Only acenaphthene accumulation was found to be greater in females. Authorities must take action if H. erythrogramma is to be currently put on New South Wales market and more biomonitoring (abundance and toxicants levels of both urchin and abiotic) must be done in different locations along the coast before the species is declared safe to eat and fish.
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. Chapter 6
6. General Discussion & Conclusion: Summary and
Recommendations
Summary
This study represents the initial establishment of a biomonitoring program in Sidon using the sea urchin Paracentrotus lividus. There was generally a low abundance of sea urchins throughout the five years of the study (2005-2009). The exception was 2007, when following a sea blockade imposed along the Lebanese coast, a partial stock recovery was observed. This suggests that the major drive of urchin abundance in the area is harvesting. When the abundance and the biomass of urchins were compared to other areas of the Mediterranean, they were found to be in the lowest range. Only the size of the sea urchins was similar to the mean size of urchins in other parts of the Mediterranean Sea. Distance from a major polluting source (waste deposit mountain) did not have any effect on the abundance, biomass and size nor did it have any impact on the amount of the metals (nickel, vanadium, lead and cadmium) concentrated in the different body parts of the urchins. Body parts accumulated metals to different extents. It is believed that in addition to the waste deposit mountain, other sources of industrial and domestic wastes, are being directly discharged in the area. This may be contributing to a more general distribution of contamination in the vicinity of Sidon‘s harbour and may have obscured any clear gradient of contamination from the waste mountain. This contamination varied between years and sea urchins, collected in 2008, had higher concentrations of metals than those of 2007. Again, this may be related to the war, which shut down much of the local industry during 2006-2007.
Given the extent of the contamination it might be argued that pristine reference locations would make useful comparisons with Sidon. However, the choice of pristine locations is not logistically simple since the extent of the Lebanese coast (225 km) is subject to the release of untreated effluent (World et al. 2003, UNEP 2007). In 2007, I chose a location nominated as a marine protected area (Tripoli North of Lebanon), and I did abundance and biomass censuses as well as analyses of metal concentration in different body parts of P. lividus found in that area. I found that the abundance and biomass of urchins in
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Tripoli were greater than that of Sidon‘s coast suggesting the effectiveness of the ―Marine Protected Area‖ classification. However, the metal concentrations in sea urchins of Tripoli were much greater than the concentrations found is urchins of Sidon. It is clear that Marine Protected Areas are only protected from harvesting and may be subject to the impact of many other anthropogenic activities. Along the coast of Tripoli, many different sources of contamination discharge into the protected area and since there is a fishing ban on the sea urchins, these urchins may have longer time to accumulate metals and other toxicants. Fortunately, the urchin population appears to be able to sustain itself in the presence of high contamination loads. Further research is necessary to test for toxic effects of metal loads on the biology of this species.
I recommend sustaining the monitoring program around Sidon but extending it to a range of other locations along the Lebanese coast. Only by keeping this program, will we be able to observe and predict toxicant concentrations in wild urchins that are subject to harvesting and human consumption. It is very important for the Lebanese authorities and the MedPol to assess the situation/outcome as presented by this study especially since a) the sea urchins’ roe content of lead has already exceeded the recommended food safety regulation b) the Lebanese are eating urchins without being informed of the potential toxicity of the lead and c) the quality of the water around Sidon is bad with potential implications for local biota. The effect of the high lead levels on the health of the organism itself should be taken into consideration due to the importance of the urchin in the food chain. In continuing and expanding the monitoring program, Lebanon would be fulfilling the regulations of international protocols. It is also recommended to expand this monitoring to include organic toxicants and radio nuclides since the latter are now being regulated (FAO/WHO 2009). It is also necessary to seek a viable alternative solution to the storage of waste at the very edge of the coast where it is likely to end up on the sea floor following storms or municipal activity. It is also recommended that monitoring takes place in winter time when the roe of sea urchins has regressed and to look at the ceolomic fluid and calcite contents. The most important issue for the ecology of the organism may well be the management of the fishery. The Lebanese authorities must urgently establish sustainable quotas for the fishing of urchins in order to protect the species from local extinction.
The second aim was to follow the progress of metal accumulation in the test, spines and teeth of P. lividus as well as accumulation of a cocktail of toxicants (inorganic, organic and bacterial) in the roe of P. lividus. For this experiment, metals known to be present in oil spills (nickel,vanadium, lead and cadmium) and which have great effects on health of
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both marine organisms and their consumers, were used for the exposure experiments in three concentrations. Roe accumulated the most contamination followed by the test, spines and teeth in decreasing order. As for the calcites, there was a delay in metal accumulation for the first two weeks. It is possible that the sea urchin will accumulate metals in areas of the body which attempt to control and detoxify these metals (in this case the roe), not only when the toxicants levels become toxic to the roe but also when the roe binding sites are saturated and can no longer be active or bind or sequester. Only when these levels become toxic to the roe or when there is a saturation of metallothionein, will the physiological response result in metals stored in the calcites (test) or the (spines). The presence of metals like the lead and cadmium in the calcites might weaken the shell and render the organism weak and prone to predators. Some metals like nickel favour the calcites regardless of whether the roe is mature or not. Exposure to higher concentrations of metals did not always result in greater accumulation efficiency perhaps due to saturation of binding sites or because the organisms were under stress and dying. The organism may use spines as way to shed metal toxicants. It was also interesting to see that faeces, algae and ceolomic fluid all had accumulated the toxicants.
With regards to the exposure to organics, the roe had reached a plateau when accumulating OCPs (dieldrin and 4,4‘DDT which metabolized into DDE in the roe). Unfortunately due to a poor recovery during the analyses of PAHs in trial 1 the fate of the low levels of acenaphthene and pyrene could not be followed. However the exposure of P.lividus to higher levels of PAHs (trial 2, 1 mg l-1 to simulate oil spill) over two weeks demonstrated that the sea urchins is indeed a good bioaccumulator of organic toxicants. However, in this trial the sea urchins were stressed and they lost their spines starting from the third day and the majority died by the end of the third week. Further research that refines the exposure method for bioaccumulation studies and potentially includes the tracking of radio-labelled trace elements would assist in our understanding of bioaccumulation kinetics. The mechanisms of accumulation and detoxification inside the roe of sea urchins are not well understood. It was one of my initial goals to use radio- labelled toxicants to follow the progress of uptake and to determine binding sites but unfortunately the infrastructure for such studies does not currently exist at the marine laboratory at AUB and permission to use it for future research at the medical facility at AUB is currently under review.
It was interesting to see how results from the laboratory exposure studies did not always match the results from the field. In Sidon, the roe was not always the organ that accumulated the most metals. This is due perhaps to the fact that the laboratory exposures
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happened in the spawning season where the roe is at its maximum maturity whereas the field sea urchins would have accumulated in other body parts during times where the roe was not mature. Moreover, the bioaccumulation was mostly through water as sea urchins were only fed briefly (12 h of food exposure over 168 total h of water exposure). That is why I was able to determine the bioconcentration factor for toxicants in the different body parts and for the two levels of exposure. BCF is one measure of toxicity and is still required by some regulating bodies like the EU. While there are some potential causes of variance in BCF they are considered to give an estimation of relative toxicity and certainly this study shows that the organic contaminant are rapidly bioconcentrated in the roe which is an issue of concern. The high levels of BCF for OCPs and for lead at the fourth week after bioaccumulation has exceeded the recommended levels set by EU (BCF less than 500 for metals) and the Stockholm convention for organics toxicants (BCF less than 5000).
I chose to use a cocktail of toxicants rather than expose urchins to individual toxicants. I considered that I would benefit from mimicking the cocktail of toxicants present in the environment so that any synergistic, additive or antagonistic effects are not missed even if these effects were not explored or measured. It would be beneficial if some side experiments are set whereby synergistic, antagonistic and additive effects are measured between two toxicants such as DDT and cadmium for example and follow its effect on the resistance of the sea urchin test.
Another beneficial outcome would probably have been reached if modern cellular approaches were taken to study bioaccumulation, however this step could not be done before establishing the fundamental step first: confirmation then determination of bioaccumulation of the used toxicants by the organism. Since I wanted to compare the bioaccumulation of the metals in different body parts, it was not enough to test toxicants in the coelomic fluid only in order not to sacrifice the animal; first because as seen in the study, there was a significant difference between body parts for different metals including the coelomic, second because injecting a needle in the sea urchin might lead to its death in 30% of cases (Luis et al. 2005).
For the bacteriological contamination, it was not clear why the amount of bacteria did not increase or fluctuate knowing that the sea urchins were exposed to total and fecal coliforms in the laboratory. It has been documented that the sea urchins can block the bacterial activity through its immune system (coelomocytes in the coelomic fluid) (Stabili et al. 1996). Effects between other toxicants and bacteria cannot be excluded.
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The bioaccumulation and depuration experiments for the P. lividus were always done between the months of May and September because it is the spawning season and the roe is usually at its maximum stage of maturity as reflected by the gonad index (Dincer et al. 2007a, Sellem et al. 2007, Garmendia et al. 2010). The roe can shrink (thin) in wintertime and becomes hence unsuitable for roe studies.
It is recommended that more experiments be done for individual radio labelled toxicant scanning any possible metabolites produced by the roe via chromatography testing using GC-MS. These kinds of experiments can bring more understanding of the accumulation mechanism and bridge the gap between exposure analysis and affects at the cellular level. It would be interesting to conduct exposure studies using 4,4’ DDT and cadmium to see if there is any singular or additive effect on calcites. Also if exposure periods can be extended to examine, if after the delay in uptake observed in the shell, a linear accumulation may be observed. It is also recommended to calculate BCF in studies where longer exposures are used at lower concentrations to see if steady state can be reached. Biomagnification through food web is worth exploring in sea urchins using clean water and algae previously contaminated with organic toxicants such as OCPs and PAHs.
The third aim was to attempt depuration in P. lividus after feeding the sea urchins toxicants in levels estimated to be present in their natural habitat following oil spill incident similar to the one that occurred in Lebanon in 2006. It also attempted to determine the kinetic rate for depuration and Biological Half-Life to be used by the aquaculture industry. The translocation experiment was more successful in some aspects than the laboratory depuration (assisted with probably an insufficient amount of EDTA). The fluctuations encountered in the laboratory experiment might be due to some mobilisation from the coelomic fluid to the roe during the sudden exposure in our laboratory. The chronic exposure will probably not show such fluctuation when depurating (e.g. as in my final Australian experiment) because the roe and the coelomic fluid would have had more time to reach a certain equilibrium among them and the coelomic fluid would have slowly but more efficiently exchanged the toxicants with its surrounding media. The sudden drop in the levels of toxicants might stem from the way they were stored and the fact that the toxicants were acquired in sudden exposure mode. Slower depuration might also mean the sea urchins were chronically exposed. P. lividus bioaccumulated and depurated lead, dieldrin and 4,4‘DDE from its roe. While vanadium was neither taken up nor depurated, nickel was taken up in small amounts but not depurated. Cadmium was taken up by the urchin but was not depurated. This is a problem for the aquaculture industry (for whom the BHL were established), the managers of
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fisheries, and the consumers especially, if more than 14 days are needed to depurate cadmium (Warnau et al. 1995 noted 1 to 2 years). No depuration was done for the other body parts of the sea urchins even if these parts were to be used in pharmaceuticals, because toxicants can be removed from these parts while they are processed industrially. For example, in the case of cadmium in calcium tablets, it might imply more treatment with complexing agents to reach the final product so that it can be in compliance with the stringent regulations imposing more financial burden on the consumer. But the roe cannot be treated for cadmium (like it is done for bacteria) in the same way as its test is treated for pharmaceuticals before being consumed
Continuous monitoring of locations from which sea urchins are commercially fished is important and recommended for the sake of both the consumers and the marine organism itself. Experiments should be set up to depurate chronically exposed animals after determining their toxicant content. In doing so, all scenarios can be covered and BHL become more incorporated so that aquaculture industry can benefit from outcomes. It is interesting to do some exposure experiments (chronic and sudden) followed by depuration and test the progress of bioaccumulation and depuration in the roe and coelomic fluid to determine if mobilisation is indeed occurring between these two body parts. It would also be worth doing some radionuclide accumulation and depuration exposure in sea urchins in light of recent food and water radioactive contamination in Japan
Heliocidaris erythrogramma in Fairlight Australia did not fare better than the P. lividus in Sidon Lebanon. The lead‘s level in H. Erythrogramma‘s roe also exceeded the WHO and the Australian guidelines. The roe of organisms collected from the field had the greatest concentrations of metals followed in decreasing order by the test, spines and teeth which echoes what was witnessed in the laboratory exposure experiments and this may suggest that the urchins are chronically exposed to toxicants in Fairlight. H. Erythrogramma is also a good bioaccumulator of metals and PAHs. Metals specifically are transmitted to eggs in levels that are yet to be studied to see if it compromises the success of fertilisation and hatching. The differences in accumulation between genders of H erythrogramma showed that it is metal specific. Females accumulated more lead than males, while males accumulated more cadmium and nickel than females. Unlike other metals, no gender difference in bioaccumulation was observed for vanadium . In PAHs, after adjusting the lipid content for both genders, males accumulated more fluoranthene and pyrene while females accumulated more acenaphthene.
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H erythrogramma was collected between November and December because the first two weeks of December are the documented spawning time for this organism (Laegdsgaard et al. 1991). Any deviation from this period can spoil any fertilisation experiments even if it was attempted by artificial means.
I recommend monitoring of the Fairlight area not only for metal content but for other organic contaminant known to exist in the field such as dioxins. Abundance and biomass must be monitored as well. Looking at the effect of high lead on the fertilisation and hatching success of H erythrogramma is worth exploring to predict the effect of this on the future of this species especially if a fisheries is established in the region. Comparing metals levels of pre and post spawning sea urchins is interesting to see how much of the metal concentration was purged and follow it up with some fertilisation experiment.
There could be some argument as to why the sea urchin was sacrificed in all experiments instead of perhaps attempting to simply test its spines or the coelomic fluid. There are at least two reasons for not taking the coelomic fluid. First because the bioaccumulation occurs in most body parts and the coelomic fluid might not reflect the real picture, especially, that toxicants might accumulate more in the roe. Second, taking the coelomic fluid can only happen if a syringe was used a fact, which by itself, can kill the organism within hours (Luis et al. 2005). The easiest example lies with the spawning/fertilisation experiments or any other experiments where it is imperative to know the sex of the organism before commencement of any trial. In this kind of experiments, the organism is injected with potassium chloride to induce spawning and determine the sex of the organism. If it was easy and feasible to simply take a biopsy from the roe, all sexes would have been determined that way and standard operating procedure would have been put in place. So the mere fact of doing the injection kills the animal, which makes additional experiments impossible. As for the spines, it is not representative to take only the spines of the organism as a measure to evade scarifying it, as the spines can be shed by organisms in time of stress ( valuable information could be lost) and spines regrow in the same organism. Hence, testing the spines alone can become a liability in the biomonitoring process.
Comparison of two species
Both urchins are considered good bioaccumulators of some organic, inorganic toxicants as well as bacteria. However, it cannot be said that both species have fulfilled the general requirements of a biomonitor. Although they are sedentary and ubiquitous; nevertheless:1- They do not bioaccumulate all toxicants within a short defined period.2-
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they do not show straightforward relationships between the concentrations of toxicants in the medium (some organic toxicants like OCPs ) and the concentrations of toxicants in the roe. 3-They shed their spines in times of stress and hence not all their body parts can be used for biomonitoring without sacrificing the organism. 4- They depurate but the rate depends on the type of the exposure (whether sudden or chronic), on the season of uptake (spawning) and on the concentrations and specificity of toxicants. Nevertheless, it is important to understand bioaccumulation kinetics within urchins because they represent a food source for humans.
Therefore some lessons and conclusions could be drawn which will help determine in which scenario sea urchins can or should be used as biomonitors: 1-When the exposure is chronic rather than transient. This is because the uptake during a transient exposure is rapid but so is the depuration which is then divided into two phases a quick sharp decline and a more slower decrease till steady state (as was the case with P. lividus). So in case of an oil spill, the load of toxicants in body parts especially in roe might decrease with time and not reliably show the exposure if not sampled immediately and if seawater medium toxic concentrations are back to normal background levels. When the exposure is chronic, uptake was slow the depuration was also slow to non-existent (e.g. H. erythrogramma). In this case the roe may be considered a useful biomonitor 2-In general, both species have tendency to accumulate more contamination in the roe than any other body part tested. So it is wise to fully study the goals and timing of the biomonitoring. It is reasonable to say that body parts could be used all year round except for the roe which can only be harnessed during spawning season (the time when the roe is at its maximum density) which is a limitation in biomonitoring. When using sea urchins collected outside spawning season, the roe of P. lividus and H. erythrogramma cannot be used for organic or metal bioconcentration determination, because the roe would have shrunk to a thin layer and lipids would be at their lowest reserve - a fact which will limit organics uptake (Laegdsgaard et al. 1991). Among the calcites, spines are not a reliable tool as they can be shed in times of stress. Instead of the roe and spines, perhaps the coelomic fluid could be considered as a short term bioindicator and the skeleton for mid- term biomonitoring especially in seasons outside spawning.
P. lividus has been well studied in ecotoxicology while H erythrogramma has not been subject to toxicity or bioaccumulation tests prior to this work. The findings in my thesis have echoed findings from other researchers in general regarding P. lividus (Coteur et al., Portocali et al., Danis et al., and Auernheimer et al.) and Warnau et al. (2005) in particular, especially as far as cadmium behaviour is concerned (whether in uptake or
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depuration). Unfortunately, the same cannot be said about H. erythrogramma, as no studies of metal accumulation were done on this species, which makes this study the first (to the best of my knowledge) to set some figures for metals uptake and depuration. For the organic toxicants, I did not expect to find equal ratios of BCF in both species first because of their different fat percentage (The lipid content of mixed gender of H erythrogramma is 25% whereas it is 14% for the P. lividus.) and second because they are two different species coming from different environments.
The abundance (witnessed in the field) and biomass of H erythrogramma appeared higher than those in the P. lividus which suggests that the H erythrogramma is not heavily fished and its population is more stable than the population of P lividus in Sidon.
The sizes of both sea urchins species were similar even though one population is more disturbed than the other. Recent studies have shown that the sea urchin is investing more in its reproduction than in its somatic growth. But recent findings has also shown that the climate change is resulting in smaller shape for benthic organism plus some difficulties for reproduction (Byrne et al. 2010). The findings of this study and the climate study (Byrne et al. 2010) suggest that the future is not so bright for both sea urchins and humans. Whether they are in the Mediterranean or in Australia, whether it is P lividus or H. erythrogramma the future is dim for these benthic organisms.
P lividus and H. erythrogramma will be overlooked most probably by public health sector too given that their consumption is low relatively to fish and other bivalves. Indeed food safety regulations take into account the daily/weekly intake of the contaminated food to establish standards and to interfere in case limits are exceeded. Sea urchins are not high on that priority list. However, the trigger is set now and action needs to be taken to remedy this situation. Both urchins suffer from high lead levels which exceeds safety limits. These lead levels will be biomagnified subsequently into higher trophic species such as fish. Fish, in turns, are highly consumed by humans and this fact alone must prompt fishery management and health authorities to act on the remediation solutions.
Conclusion
Sea urchins have proved that they are potentially important biological indicators of ecological disturbance only but good bioaccumulators of inorganic and organic contaminants. Paracentrotus lividus abundance in the Mediterranean Sea is affected by harvesting activities but in some areas may also be affected by contamination of their habitat. Comparing the sea urchins of Lebanon to those of the rest of the Mediterranean
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basin, showed that sea urchins of Lebanon were in the lowest range of abundance and biomass, had similar sizes to the rest of the Mediterranean but contained higher metal loads. In Sidon, where sea urchins are fished and consumed by people, Paracentrotus lividus had loads of lead much greater than the food safety guidelines. Lebanese authorities must take action for the sake of human health and for the survival of the biota. Among all the toxicants tested for bioaccumulation and depuration, cadmium was the only toxicant which was not eliminated by the urchin. Therefore, if sea urchins keep accumulating cadmium, this may have a devastating effect on both the organism and the fishery. First, it will eventually reach a level which is above food safety regulations. Second, it will affect the health of both the organism and the consumer and finally, it will not be able to depurate the cadmium in relatively small amount of time to save both its reproduction and the fishery. The roe of the sea urchin is an important site where toxicants are taken up in higher levels than the calcites. The roe is also the site that has the ability to metabolize organic compounds. In fisheries, the roe is the edible part, so caution must be exercised before consumption of this organ and before establishing new fisheries. In Australia, monitoring abundance and baseline levels of Heliocidaris erythrogramma is important before starting a new fishery. H. erythrogramma has already exceeded international standard for food safety limits as far as lead concentrations are concerned. Effects on reproduction are of major concern since eggs shed after spawning had accumulated a substantial concentration of metals.
This study will hopefully be beneficial to many stakeholders: The Mediterranean Pollution program and the people of the Mediterranean (Lebanese included), the Lebanese authorities, the aquaculture industries in the Mediterranean and in Australia, Australian authorities and to the scientific community in general. It shed light on the ability of sea urchins to accumulate organic toxicants, to the best of my knowledge, not explored previously in the P lividus species. It has established depuration rates for many of the toxicants studied in the exposure experiment, shed light on the probable physiological defence tools of sea urchins (metabolites produced by the roe, shedding spines) and determined background levels of sea urchins native to both the Lebanese (P. lividus) and the Australian (H. erythrogramma) coast. To the best of my knowledge, H. erythrogramma has never been tested for metal content or ability to take up metals or PAHs.
Sea urchins still have lots of secrets to reveal, especially in relation to the functioning of their immune system. Since, sea urchins and humans share more than 7,000 genes, it is believed that by mapping sea urchin DNA, scientists may uncover information about how
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human diseases develop. Subsequently, understanding how urchins respond to contaminant exposure may provide insights into human health. The sea urchins fate and that of humans‘ seem to be intertwined and their health is a reflection of ours. So monitoring urchins, depurating them and performing further studies on them can only bring advances to both species.
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Appendices
Appendix 1- Abbreviations
172
Appendices Abbreviations Full name/s ANOVA Analysis of Variance µg Microgram µl Microliter ACS Chemical Abstract Service AES Atomic Emission Spectrometry AMSA Australiam Maritime and Safety Authority ANZECC Australia and New Zealand Environment Council Committee AOAC Association of Official Analytical Chemists APHA American Public Health Association Alltech Heliflex capillary columns for mass spectrometers 5% phenyl, 95% AT-5MS dimethylpolysiloxane ATSDR Agency for Toxic Substances and Disease Registry AUB American University of Beirut BCF Bioconcentration Factor BHL Biological Half Life C Contaminated ca Circa, around, in approximately; about CAP College of American Pathologists Cd Cadmium cm Centimeter d Day DCM DiChloroMethane DDE Dichloro-Diphenyl-Dichloroethylenene DDT Dichloro-Diphenyl-Trichloroethane df Degree of Freedom DI De Ionized DNA Deoxyribonucleic acid E.coli Escherichia coli e.g. Exempli gratia, For example EDTA Ethylene DiamineTetra Acetic acid EIM Electron Impact Mode EPA-USA Environment Protection Agency-United States of Ameri EU European Union EVL Environment Core Laboratory F Fished FAO Food and Agricultural Organisation FDA Food and Drug Agency g Gram GC-ECD Gas Chromatography- Electron Capture Detector GC-MS Gas Chromatography- Mass Spectrometry GF-AAS Graphite Furance-Atomic Absorption- Atomic Absorption Spectrometry h Hour/s HCH Hexa-Chloro-CycloHexane HCL Hydro Chloric Acid HPLC High Performance Liquid Chromatography IARC The International Agency for Research on Cancer ICP-MS Ion Coupled Plasma-Mass Spectrometry IMO International Maritime Organization IS Internal Standard ISO International Standard Operation IVF In Vitro Fertilisation JW-DB-5MS Agilent JW Scientific DB-5ms GC Columns5% phenyl, 95% dimethylpolysiloxane k Kinetic rate
173 Appendices Abbreviation Full name/s KCl Potassium Chloride Km Kilometer Kow Octanol-Water Partition Coefficient l Liter m Meter m2 square meter MAP Mediterranean Action Plan MedPol Mediterranean Pollution METAP Mediterranean European Technical Assistance Programme ms Millisecond MS Mean Squares mS/cm Milli Siemens per Centimeter MSD Mass-Septrometer Detector MW MolecularWweight N/A Not Available nb Number NGO Non Governmental Organisation NH3-N/L Nitrogen bound as Ammonia per Liter Ni Nickel NIST National Institute of Standards and Technology NM Not Mentioned NSW New South Wales NTU Nephelometric Turbidity Units oC Degree Celsius OCP Organo-Chlorinated Pesticide/s OQVP Operational Qualification and Performance Verification P Protected PAH Polycyclic Aromatic Hydrocarbon/s Pb Lead PCB Poly Chlorinated Biphenyls PERMANOVA Permutational Analysis of Variance POP Persistant Organic Pollutants PSA Primary Secondary Amines PTFE Poly Tetra Fluoro Ethylene PVC Poly Vinyl Chloride QC Quality Control QuEChERS Quick Easy Cheap Effective and Safe R Reference REMPEC The Regional Marine Pollution Emergency Response Centre for the Mediterrranean Sea ROS Reactive Oxygen Species Scuba Self Contained Underwater Breathing Apparatus SE Standard Error SHOROC Shore Regional Organisation of Councils SIM Selective Ion Monitoring SNK StudentNewman-Keuls SPSS Statistical Package for the Social Sciences, Software TK-TD Toxico Kinetics-Toxico Dynamics UNEP United Nations Environment Program UNFPA United Nations Population Fund Agency USA United States of America V Vanadium Wt Weight Wet weight WHA World Health Assembly WHO World Health Organisation
174 Appendices Appendix 2 Body parts and taxonomy of sea urchins.
Body parts of the sea urchin
Spines Teeth
Test
Roe
Taxonomy
Paracentrotus lividus
Kingdom Animalia Phylum Echinodermata Class Echinoidea (Sea Urchins, Sand Dollars, Heart Urchins) Family Echinidae Paracentrotus lividus - Sea Urchin
Heliocidaris erythrogramma
Kingdom Animalia Phylum Echinodermata Class Echinoidea (Sea Urchins, Sand Dollars, Heart Urchins) Family Toxopneustidae Heliocidaris erythrogramma - Purple Sea Urchin
175