Climate Change and Marine and Freshwater Toxins 9783110333596, 9783110333039

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Botana, Louzao, Vilariño (Eds.) Climate Change and Marine and Freshwater Toxins

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Botanica Marina Dring, Matthew (Editor-in-Chief) ISSN 0006-8055, e-ISSN 1437-4323

Climate Change and Marine and Freshwater Toxins

| Edited by Luis M. Botana, M. Carmen Louzao and Natalia Vilariño

Editors Prof. Luis M. Botana Universidad de Santiago de Compostela Facultad de Veterinaria Departamento de Farmacología 27002 Lugo, Spain [email protected] Prof. M. Carmen Louzao Universidad de Santiago de Compostela Facultad de Veterinaria Departamento de Farmacología 27002 Lugo, Spain [email protected] Prof. Natalia Vilariño Universidad de Santiago de Compostela Facultad de Veterinaria Departamento de Farmacología 27002 Lugo, Spain [email protected]

ISBN 978-3-11-033303-9 e-ISBN (PDF) 978-3-11-033359-6 e-ISBN (EPUB) 978-3-11-038261-7 Set-ISBN 978-3-11-033360-2 Library of Congress Cataloging-in-Publication Data A CIP catalog record for this book has been applied for at the Library of Congress. Bibliographic information published by the Deutsche Nationalbibliothek The Deutsche Nationalbibliothek lists this publication in the Deutsche Nationalbibliografie; detailed bibliographic data are available on the Internet at http://dnb.dnb.de. © 2015 Walter de Gruyter GmbH, Berlin/Boston Cover image: © Dennis Kunkel Microscopy, Inc./Visuals Unlimited/Corbis Typesetting: PTP-Berlin, Protago-TEX-Production GmbH, Berlin Printing and binding: CPI books GmbH, Leck ♾ Printed on acid-free paper Printed in Germany www.degruyter.com

Preface Climate change and its far-reaching impacts compel us to question the value we give to nature and the human relationship with it. Water touches many subject areas that are important in our daily lives. At some point it influences what we eat and drink, and how we access it. Therefore, we decided to edit a book on climate change and water toxins that includes climate trends and effects, physicochemical measurements, water quality parameters, marine and freshwater toxins, toxin detection, phytoplankton and zooplankton, invertebrates and fish. There is no historical record to compare the amounts of toxins existent now and a century or more ago. Toxins are identifiable as a result of modern science, and thereby their presence, structure or levels in food have only been known for a short time. The use of mass spectrometers is rather recent, and the existence of certified standards only goes back a few years. Therefore, it is very complex to establish a solid link, using the scientific method, between climate change and toxins. But it is clear that something is happening – not only because modern technology allows us to track the changes easily, but also because the trend is that more and different toxins are appearing in new locations and products. Although climate change is frequently related to extreme weather episodes and rising sea levels in the media, a lesser known fact is that new toxins will appear in areas and products where they presently do not occur. Despite the fact that scientific evidence may not always be available to prove or disprove perceived potential harms of climate change and their links with toxins, this book offers quantitative compelling evidence of the many complex interactions that must be considered from primary toxin producers up the food chain to humans. In the case of marine toxins, although ballast water, international trade, etc. may be a source of new intoxications and blooms, it is very clear that some regions are hotspots for many compounds. Likewise, eutrophication of lakes is a source of cyanobacterial blooms. The US had never had a diarrheic episode until Texas witnessed one a few years ago. Europe had never had a tetrodotoxin intoxication from shellfish until a few years ago, ciguatoxin intoxications are becoming frequent after ingestion of fish from the Southern European Atlantic Ocean, and aerosols with ostreocin from Gambierdiscus are now a problem in Mediterranean beaches every year. A similar problem is being observed in freshwater, as the expansion of cyanobacteria and their toxins has become a worldwide problem; this adds to the deleterious effect of human pollution in drinking water. Something is happening that was not previously reported and may be explained by increased water temperatures in both lakes and seas. This book intends to cover the main aspects of the possible relation between climate change and freshwater and marine toxins: prediction models and management of harmful algal blooms; influence on food security and food production; legislation; drinking water and cyanobacteria blooms; and sex change in toxin vectors. This last

VI | Preface

topic, sex change, serves as an introduction to a new area of research – the role of climate change in basic physiological processes. Very little information is currently available on this subject. This book has brought together a group of international experts. Contributing authors expand the framework of possibilities for appropriate assessment of climate change impacts on marine and freshwater toxins, which in turn directly impacts the natural environment, human health and sustainability. The book is an excellent introduction to this complex topic or a useful supplement to courses in the field of ecotoxicology. In short, it is a must-read book for all who are interested in toxins and how climatic conditions can modify them – from the general public or students to toxicologists, food technologists, pharmacologists, analytical chemists, ecologists, biologists, veterinarians and physicians. Last, and by no means least, we wish to thank all the authors. They were not only very generous with their time, they were also bold enough to commit to write a chapter on an especially difficult topic and use their prestigious names in their chapters. For this, we are greatly thankful to all of them. We hope the book helps in understanding the potential risks caused by climate in one particularly sensitive area: food and drinking water.

Contents Preface | V List of contributing authors | XV Josefino C. Comiso 1 Variability and trends of global sea ice cover and sea level: effects on physicochemical parameters | 1 1.1 Introduction | 1 1.2 Variability and trends of global sea ice | 2 1.2.1 Arctic Region | 5 1.2.2 Antarctic Region | 8 1.3 Variability and trends in sea level | 12 1.3.1 Contributions from warming oceans | 13 1.3.2 Contributions from glaciers, ice sheets and others | 15 1.4 Effects on physicochemical parameters | 19 1.4.1 Large-scale changes in surface temperature | 19 1.4.2 Large-scale changes in plankton concentration and primary productivity | 20 1.4.3 Changes in other physicochemical parameters | 26 1.5 Discussion and conclusions | 29 Begoña Espiña, Marta Prado, Stephanie Vial, Verónica C. Martins, José Rivas, and Paulo P. Freitas 2 New techniques in environment monitoring | 35 2.1 Introduction | 35 2.2 In situ harmful algal bloom monitoring | 36 2.2.1 Optical remote sensing | 36 2.2.2 Automated monitoring | 38 2.2.3 HABs sampling based on absorption | 42 2.3 Liquid chromatography and mass spectrometry | 44 2.4 Biosensors for HABs monitoring | 46 2.4.1 Optical biosensors | 49 2.4.2 Electrochemical biosensors | 51 2.4.3 Mass biosensors | 51 2.4.4 Magnetic-based biosensors | 52 2.5 Advances in nanotechnology for HAB detection | 53 2.5.1 Nanoparticles | 54 2.5.2 Analytical nano-applications | 55 2.6 Molecular biology-based techniques for HABs detection | 64

VIII | Contents

2.6.1 2.6.2 2.6.3 2.6.4 2.6.5 2.7

Overview | 64 DNA/RNA targets | 65 Hybridization-based techniques | 70 Amplification-based techniques | 72 Aptamers for toxin detection | 75 Future perspectives | 76

Mikko Nikinmaa and Katja Anttila 3 Responses of marine animals to ocean acidification | 99 3.1 Introduction | 99 3.2 What causes ocean acidification | 99 3.2.1 Effect of atmospheric carbon dioxide loading | 100 3.2.2 Influence of primary production | 101 3.2.3 Carbon balance in coastal areas | 101 3.2.4 Interactions between temperature changes and ocean acidification | 102 3.3 Processes of animals that are expected to be affected | 102 3.3.1 pH regulation | 102 3.3.2 Calcification | 107 3.3.3 Development | 108 3.3.4 Oxygen transport and metabolism | 110 3.3.5 Behavior | 114 3.4 Conclusions | 115 Shauna Murray, Uwe John, and Anke Kremp 4 Alexandrium spp.: genetic and ecological factors influencing saxitoxin production and proliferation | 125 4.1 Introduction | 125 4.2 Alexandrium taxonomy, phylogenetics and species evolution | 126 4.3 What are saxitoxins? | 129 4.3.1 Which species produce saxitoxins? | 130 4.3.2 The sxt genes in dinoflagellates | 131 4.4 Ecological factors influencing Alexandrium spp. proliferation and toxicity | 133 4.4.1 The role of ecophysiological adaptations in ecology and bloom formation of Alexandrium life cycles | 133 4.4.2 Mixotrophic nutrition | 133 4.4.3 Allelopathy | 134 4.5 Effects of environmental factors on Alexandrium proliferation and toxicity | 135 4.5.1 Nutrients | 135 4.5.2 Temperature | 135

Contents | IX

4.5.3 4.5.4 4.6

CO2 | 138 Salinity | 139 Adaptation to changing climate conditions | 141

Susanna A. Wood, Jonathan Puddick, Hugo Borges, Daniel R. Dietrich, and David P. Hamilton 5 Potential effects of climate change on cyanobacterial toxin production | 155 5.1 Introduction | 155 5.1.1 Microcystins and nodularins | 156 5.1.2 Cylindrospermopsins | 157 5.1.3 Saxitoxins | 157 5.1.4 Anatoxin-a and homo-anatoxin-a | 157 5.1.5 Anatoxin-a(S) | 158 5.1.6 Lipopolysaccharides (LPS) | 158 5.2 Effects of climate change on common toxin producing species | 159 5.2.1 Microcystis | 160 5.2.2 Cylindrospermopsis | 161 5.2.3 Dolichospermum | 161 5.2.4 Planktothrix | 162 5.2.5 Phormidium | 163 5.3 Effects of climate change on toxin regulation | 164 5.3.1 Microcystins | 164 5.3.2 Nodularins | 166 5.3.3 Cylindrospermopsins | 166 5.3.4 Saxitoxins | 167 5.3.5 Anatoxins | 167 5.4 Climate change and its effect on cyanobacteria and toxin production in Polar environments | 168 5.5 Conclusions | 170 Gustaaf M.Hallegraeff 6 Harmful marine algal blooms and climate change: progress on a formidable predictive challenge | 181 6.1 Introduction | 181 6.2 Algal bloom range extensions and climate change | 182 6.3 Range extensions further aided by ship ballast water transport | 184 6.4 The formidable challenge of predicting phytoplankton community responses | 187 6.5 We can learn from the fossil record, long-term plankton records and decadal scale climate events | 188 6.6 Mitigation of the likely impact on seafood safety | 188

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Elke S. Reichwaldt, Som Cit Sinang, and Anas Ghadouani 7 Global warming, climate patterns and toxic cyanobacteria | 195 7.1 Introduction | 195 7.2 The effect of global warming on inland water bodies | 196 7.2.1 Direct effects of global warming on inland water bodies | 196 7.2.2 Indirect effects of global warming on inland water bodies | 197 7.3 The ecology of cyanobacteria and toxin production | 203 7.3.1 Environmental factors affecting cyanobacterial biomass | 203 7.3.2 Environmental factors affecting microcystin production | 204 7.3.3 Ecological factors affecting cyanobacterial blooms: competition | 206 7.4 Direct and indirect effects of global warming on cyanobacterial growth | 208 7.4.1 Temperature, stratification, and mixing | 215 7.4.2 Nutrients | 216 7.4.3 Salinity | 217 7.4.4 Turbidity and pH | 217 7.5 Direct and indirect effects of global warming on microcystin concentration | 217 7.6 Why should we care? | 219 Aristidis Vlamis and Panagiota Katikou 8 Human impact in Mediterranean coastal ecosystems and climate change: emerging toxins | 239 8.1 Introduction | 239 8.2 Mediterranean coastal ecosystems | 240 8.2.1 Human impact | 242 8.2.2 Socio-economical implications of Climate Change | 244 8.2.3 Effect to ecosystem from extreme events of climate change | 245 8.2.4 Ecological response to Climate Change | 246 8.3 Emerging toxins in the Mediterranean Sea | 248 8.3.1 Identified emerging toxins and climate change effects | 249 8.4 Conclusion | 259 Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray 9 Gambierdiscus, the cause of ciguatera fish poisoning: an increased human health threat influenced by climate change | 273 9.1 The genus Gambierdiscus | 273 9.2 Morphology and phylogenetics | 274 9.3 Geographic distribution and abundance | 279 9.3.1 The Pacific and Indian Ocean Regions | 282 9.3.2 The Atlantic Ocean Region | 282 9.4 CTXs and MTXs | 283

Contents | XI

9.5 9.6 9.7

Toxicity of different species of Gambierdiscus | 288 Detection of CTXs and MTXs in seafood | 289 Conclusion | 303

Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani 10 Control and management of Harmful Algal Blooms | 313 10.1 Introduction | 313 10.2 Global water crisis | 313 10.3 Cyanobacteria and cyanotoxins | 314 10.4 Cyanobacterial prevention and mitigation | 315 10.5 Cyanobacterial management | 320 10.6 Case study: The management of cyanobacteria in waste stabilization ponds | 323 10.7 Treatment of cyanobacteria and cyanotoxins with hydrogen peroxide | 326 10.8 New techniques for the control and characterization of cyanobacterial blooms | 335 10.8.1 Allelopathic control of cyanobacteria | 335 10.8.2 Optimization of the FDA-PI method using flow cytometry to measure metabolic activity of cyanobacteria | 336 10.9 New perspectives and future directions | 338 Joaquín Espinosa, Sara Silva-Salvado, and Óscar García-Martín 11 Global climate change profile and its possible effects on the reproductive cycle, sex expression and sex change of shellfish as marine toxins vectors | 359 11.1 Introduction | 359 11.2 Shellfish as marine toxins vectors | 360 11.2.1 General considerations | 360 11.2.2 Global increase in HABs | 362 11.2.3 Global climate change | 365 11.3 Reproductive cycle, sex expression and sex change in shellfish | 378 11.3.1 Reproductive cycle, reproductive period and sex expression in bivalve mollusks | 378 11.3.2 What is sex? | 379 11.3.3 Sex determination: everything happens in the embryo | 380 11.3.4 Sex determination of the gonad and sex differentiation of primordial germ cells (PGCs): molecular basis and regulation | 381 11.3.5 Gonad somatic sex and germline sex in bivalve mollusks | 382 11.3.6 Sex, sex reversal, types of sexuality and sex change in bivalve mollusks | 384

XII | Contents

11.3.7 11.3.8 11.3.9 11.4

What does sex change mean and how could this process be performed by bivalve mollusks? | 391 Temperature, photoperiod, reproductive cycle and sex change in bivalve mollusks | 393 Climate change, reproductive cycle, sex expression and sex change in bivalve mollusks | 398 Concluding remarks | 402

M. Carmen Louzao, Natalia Vilariño, and Luis M. Botana 12 Effects on world food production and security | 417 12.1 Introduction | 417 12.2 Foodborne and waterborne diseases | 417 12.3 Zoonosis and other animal diseases | 418 12.4 Product safety in fisheries | 419 12.5 Aquaculture food production | 423 12.6 Harmful algal blooms | 423 12.6.1 Impact of temperature change on harmful algal blooms | 424 12.6.2 Acidification of waters and effect on harmful algal blooms | 426 12.6.3 Impact of sea-level rise and increased precipitation on harmful algal communities | 426 12.6.4 Microalgal toxicity | 427 12.7 Harmful algal blooms and aquatic food safety | 428 12.7.1 Predictive modeling | 433 12.8 Future perspectives | 434 Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana 13 From science to policy: dynamic adaptation of legal regulations on aquatic biotoxins | 441 13.1 Introduction | 441 13.2 Current worldwide regulations on marine phycotoxins | 441 13.2.1 Maximum permitted levels | 441 13.2.2 Official detection methods | 446 13.3 Current worldwide regulations on cyanotoxins | 447 13.4 New occurrences of toxic episodes challenge protection of consumer’s safety | 455 13.5 Limitations for the development and implementation of new regulations: from science to policy or from policy to science? | 457 13.5.1 Technical limitations for recent/future toxin regulations | 457 13.5.2 Toxicological limitations for new toxin regulations | 461 13.5.3 Economic limitations | 465 13.6 Modification of monitoring and surveillance programs | 466

Contents | XIII

13.7 13.8

Integrative example: tetrodotoxin as a biomarker of climate change | 467 Concluding remarks | 470

Index | 483

List of contributing authors Katja Anttila Department of Biology University of Turku FI-20014 Turku, Finland Chapter 3 Dani J. Barrington Aquatic Ecology and Ecosystem Studies School of Civil, Environmental and Mining Engineering The University of Western Australia 35 Stirling Highway Crawley, WA 6009, Australia and Department of Marketing Monash Business School Monash University Wellington Road Clayton, VIC 3800, Australia and International WaterCentre 333 Ann Street Brisbane, QLD 4000, Australia Chapter 10 Hugo Borges Cawthron Institute, Nelson, New Zealand and Environmental Research Institute University of Waikato Hamilton, New Zealand Chapter 5 Luis M. Botana (Ed.) Departamento de Farmacología Facultad de Veterinaria Universidad de Santiago de Compostela 27002 Lugo, Spain [email protected] Chapter 12

Liah X. Coggins Aquatic Ecology and Ecosystem Studies School of Civil, Environmental and Mining Engineering The University of Western Australia 35 Stirling Highway Crawley, WA 6009, Australia Chapter 10 Josefino C. Comiso Cryospheric Sciences Laboratory Earth Sciences Division NASA/Goddard Space Flight Center Greenbelt, MD USA 20771 [email protected] Chapter 1 Daniel R. Dietrich Human and Environmental Toxicology University of Konstanz 78464 Konstanz, Germany Chapter 5 Begoña Espiña International Iberian Nanotechnology Laboratory (INL) Avenida Mestre José Veiga 4715-330 Braga, Portugal [email protected] Chapter 2 Joaquín Espinosa Physiology Department Faculty of Pharmacy Santiago de Compostela University (USC) 15782-Santiago de Compostela, Spain [email protected] Chapter 11

XVI | List of contributing authors

Hazel Farrell Plant Functional Biology and Climate Change Cluster (C3) University of Technology Sydney Sydney, PO Box 123, Broadway New South Wales 2007, Australia [email protected] Chapter 9 María Fraga Departamento de Farmacología Facultad de Veterinaria Universidad de Santiago de Compostela 27002 Lugo, Spain Chapter 13 Paulo P. Freitas International Iberian Nanotechnology Laboratory (INL) Avenida Mestre José Veiga 4715-330 Braga, Portugal Chapter 2 Óscar García-Martín Biochemistry and Molecular Biology Department Faculty of Pharmacy Santiago de Compostela University (USC) 15782-Santiago de Compostela, Spain Chapter 11 Anas Ghadouani Aquatic Ecology and Ecosystem Studies School of Civil, Environmental and Mining Engineering The University of Western Australia 35 Stirling Highway Crawley, WA 6009, Australia [email protected] Chapter 7 and 10 Gustaaf M. Hallegraeff Institute for Marine and Antarctic Studies (IMAS) University of Tasmania Private Bag 129 Hobart, Tasmania 7001, Australia [email protected] Chapter 6

David P. Hamilton Environmental Research Institute University of Waikato Hamilton, New Zealand Chapter 5 Uwe John Alfred Wegener Institute for Polar and Marine Research Am Handelshafen 12 27570 Bremerhaven, Germany Chapter 4 Panagiota Katikou National Reference Laboratory on Marine Biotoxins Ministry of Productive Reconstruction, Environment and Energy 3A Limnou street 54627 Thessaloniki, Greece [email protected]; [email protected] Chapter 8 Gurjeet S. Kohli Plant Functional Biology and Climate Change Cluster (C3) University of Technology Sydney Sydney, PO Box 123, Broadway New South Wales 2007, Australia [email protected] Chapter 9 Anke Kremp Marine Research Centre Finnish Environment Institute 00251 Helsinki, Finland Chapter 4 M. Carmen Louzao (Ed.) Departamento de Farmacología Facultad de Veterinaria Universidad de Santiago de Compostela 27002 Lugo, Spain [email protected] Chapter 12 and 13

List of contributing authors | XVII

Verónica C. Martins International Iberian Nanotechnology Laboratory (INL) Avenida Mestre José Veiga 4715-330 Braga, Portugal Chapter 2 Shauna A. Murray Plant Functional Biology and Climate Change Cluster (C3) University of Technology Sydney Sydney, PO Box 123 Broadway New South Wales 2007, Australia [email protected] Chapter 4 and 9 Mikko Nikinmaa Department of Biology University of Turku FI-20014 Turku, Finland [email protected] Chapter 3 Marta Prado International Iberian Nanotechnology Laboratory (INL) Avenida Mestre José Veiga 4715-330 Braga, Portugal Chapter 2 Jonathan Puddick Cawthron Institute, Nelson, New Zealand Chapter 5 Elke S. Reichwaldt Aquatic Ecology and Ecosystem Studies School of Civil, Environmental and Mining Engineering The University of Western Australia 35 Stirling Highway Crawley, WA 6009, Australia [email protected] Chapter 7

José Rivas International Iberian Nanotechnology Laboratory (INL) Avenida Mestre José Veiga 4715-330 Braga, Portugal and NANOMAG Laboratory Research Technological Institute Department of Applied Physics University of Santiago de Compostela 15782 Santiago de Compostela, Spain Chapter 2 Sara Silva-Salvado Biochemistry and Molecular Biology Department Faculty of Pharmacy Santiago de Compostela University (USC) 15782-Santiago de Compostela, Spain Chapter 11 Som Cit Sinang Aquatic Ecology and Ecosystem Studies School of Civil, Environmental and Mining Engineering The University of Western Australia 35 Stirling Highway Crawley, WA 6009, Australia Present address: Faculty of Science and Mathematics Sultan Idris Education University 35900 Tanjong Malim Perak, Malaysia Chapter 7 Stephanie S. Vial 3B’s Research Group – Biomaterials, Biodegradables and Biomimetics University of Minho Headquarters of the European Institute of Excellence on Tissue Engineering and Regenerative Medicine AvePark 4806-909 Taipas, Guimarães, Portugal and ICVS/3B’s – PT Government Associate Laboratory Braga/Guimarães, Portugal Chapter 2

XVIII | List of contributing authors

Natalia Vilariño (Ed.) Departamento de Farmacología Facultad de Veterinaria Universidad de Santiago de Compostela 27002 Lugo, Spain [email protected] Chapter 12 and 13 Aristidis Vlamis National Reference Laboratory on Marine Biotoxins Ministry of Productive Reconstruction, Environment and Energy 3A Limnou street 54627 Thessaloniki, Greece and Department of Pharmacology Veterinary School University of Santiago de Compostela Lugo 27002, Spain Chapter 8

Susanna A. Wood Cawthron Institute, Nelson, New Zealand [email protected] and Environmental Research Institute University of Waikato Hamilton, New Zealand Chapter 5 Xi Xiao Ocean College Zhejiang University Hangzhou 310058, PR China and College of Environmental & Resource Science (CERS) Zhejiang University Hangzhou 310058, PR China Chapter 10

Josefino C. Comiso

1 Variability and trends of global sea ice cover and sea level: effects on physicochemical parameters 1.1 Introduction The rapid decline in the Arctic summer ice cover minimum and the acceleration of sea level rise, as reported in recent years [1–3], have gained a great deal of attention and are regarded as among the most visible signals of anthropogenic global warming. Both phenomena have been linked to climate change either directly or indirectly and are expected to cause profound changes in the physicochemical characteristics of the polar and extrapolar regions. Historically, the high latitude regions have received little interest mainly because of general inaccessibility, harsh weather conditions and the paucity of data. The advent of satellite remote sensing has completely changed this, and our ability to monitor polar regions and especially sea ice cover, other components of the cryosphere and sea level rise has been vastly improved. In particular, polar orbiting satellite data have yielded more than 3 decades of consistent and continuous global data sets at a temporal resolution of even better than twice daily. But more importantly, the data have yielded strong evidence that dramatic changes related to climate are occurring in the polar regions. The yearly Arctic summer ice cover minimum has been studied and used to indicate that the perennial ice cover has been rapidly declining – even with only approximately 22 years (i.e. 1978 to 2000) of satellite data [4]. By perennial ice, we mean the thick ice type that normally survives the spring and summer melt. It is the mainstay of the Arctic sea ice cover and is known to have been in existence for at least 1450 years [5]. Following a dramatic decline in the Arctic summer sea ice cover in 2007 [1], the extent of the perennial ice was a record low in 2012 [6] and was found to be less than half of its extents observed in the 1980s. This has led to a realization that the Arctic region has been changing fast; this has ignited many international and national research projects in the region. While a sea ice cover decline has been expected from modeling studies, the rate of decline as projected by the models is significantly less than that actually observed from satellite data [7]. This means that there are still gaps in our knowledge of the physics of the Arctic climate system. The observed decline is consistent with the amplification of global warming in the Arctic by more than 3 times as has been reported [8]. Such amplification is in part the result of a phenomenon called “ice-albedo feedback” that is associated with the decline of the summer ice cover [9]. Such warming has affected the rest of the cryosphere in the region, including the glaciers in North America and the vast ice sheet of Greenland [10]; the latter has a sea level equivalence of greater than seven meters.

2 | Josefino C. Comiso

The observed physical changes in the Southern Hemisphere is different from those in the Arctic and appear counterintuitive but not totally unexpected, considering that the impact of global warming around the globe is not uniform [11]. Sea ice cover is observed to be expanding and some cooling is observed in large areas of the Antarctic region [12, 13]. Quantitative estimates of the rate of loss of mass observed due to melting of the massive ice sheet in Antarctica, which has a sea level equivalence of more than 60 meters, are also of interest, but results from different investigators have been inconsistent [10]. In this chapter, an overview of the current state of the global sea ice cover, changes in surface temperature, related rise in sea level and associated changes in the physicochemical characteristics of the marine environment and global climate will be presented. Sea ice is already a part of the ocean and causes a negligible increase in sea water level when it melts. However, the loss of sea ice in recent years, especially in the spring and summer, has contributed to the aforementioned amplified warming in the Arctic that in turn has caused more land ice to melt and thus a higher rate of sea level rise. It will likely cause a change in the primary productivity of the region as well. The direct impacts of changes in the global sea ice cover on the physicochemical characteristics of the affected regions are expected to be profound, but specific details are basically unknown and still the subject of many modeling and observational studies. Most studies focus on the impact of climate change in general on marine ecosystems [14], but the impact on polar marine ecosystems has become the subject of strong interest because of the rapidly retreating Arctic ice cover [15]. A significant increase in sea level is expected to cause serious negative changes on the global marine ecosystem, but most studies concentrate mainly on impacts that primarily occur in coastal regions [16] where the effects are most visible. Because of rapid changes in the polar environment in recent years, this study will make use of results from updated data sets to take advantage of an extended data record to provide an interpretation of the observed phenomena that is as accurate as possible. This will thereby enhance our current understanding of the physics of the system. The use of updated data also allows for an assessment of previous forecasts’ accuracy that relied on numerical models or statistical studies.

1.2 Variability and trends of global sea ice The sea ice cover has been considered a key component of the Earth’s climate system. Because of its high albedo and good thermal insulating property, it is very effective in limiting heat and salinity fluxes between the ocean and the atmosphere. Sea ice also redistributes surface salinity and alters surface density; this causes vertical circulation that enables upper layers of the oceans to be replenished with nutrients, oxygen and other chemicals. The ice-covered regions make up three of only four regions where deep ocean convection has been observed worldwide. They have also been the primary

1 Variability and trends of global sea ice cover and sea level

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source of bottom water that is an essential component of global ocean thermohaline circulation. Latent heat and sensible heat polynyas are also formed in ice-covered regions; this causes an alteration in the physical and chemical properties of the ocean [17]. Sea ice melt in the spring and summer also causes the formation of a stable melt surface layer that is exposed to abundant sunlight and serves as a platform for efficient photosynthesis. This phenomenon makes the region a site for phytoplankton blooms and high primary productivity during the spring and summer [18]. Sea ice also covers a large fraction (about 6 %) of the global oceans and can affect shipping, fisheries and mineral exploration. Among the most important tools that have been used to study the large-scale characteristics, variability and trends of global sea ice cover are satellite passive microwave sensors. The first imaging system was the Nimbus-5/ESMR that was launched in December 1972 and was the first to reveal the true extent and variability of global sea ice cover at a temporal resolution of about 3 days. Because of the large contrast in the brightness temperature of sea ice-covered and ice-free ocean, the sensor was able to provide the extent and general characteristics of the sea ice cover [19]. However, with only one channel available at 19 GHz and horizontal polarization, the data could only provide rough estimates of the ice concentration within the pack because of varying surface emissivity associated with different ice types and snow cover conditions. Accurate, consistent and continuous monitoring of sea ice cover started with the Nimbus-7 Scanning Multichannel Microwave Radiometer (SMMR) which was launched in October 1978 and was succeeded by a series of the DMSP/Special Scanning Microwave Imager (SSM/I) that started in July 1987 and continued up to the present. The two sensors are both dual polarized, multi-frequency and conically scanning systems for consistent coverage of the surface at a resolution of about 25 km. An even more capable system called AMSR-E was launched on board the EOS/Aqua satellite in May 2002; it provides more accurate brightness temperature data at a significantly higher resolution [20]. AMSR-E provided high quality sea ice data until it suffered instrumental problems and was turned off in 2011. It was succeeded by AMSR-2, which was launched on board GCOM-W by JAXA in 2012. Analyses of time series of sea ice data from various passive microwave sensors have been reported in several publications [12, 13, 21–24]. Different techniques were used to retrieve sea ice concentrations and therefore ice extent and ice area. However, the results on seasonal and inter-annual variability of sea ice have been generally consistent [21, 22]. The trends reported for different time periods are generally different – albeit slightly – because of the relatively short record length and the large inter-annual variability of ice cover. Color-coded maps of ice concentration averages during maximum and minimum extents for each year since November 1978 are presented in Fig. 1.1. Because of large inter-annual variability in the location of the ice edges, the images show smearing at the ice margins. In the Northern Hemisphere, the sea ice cover in the Arctic Basin is confined by surrounding land areas, but the sea ice reaches the peripheral seas and

4 | Josefino C. Comiso

(a) NH minimum average with 2014 minimum contour

(b) NH maximum average with 2014 maximum contour

(c) SH minimum average with 2014 minimum contour

(d) SH maximum average with 2014 maximum contour

0%

25%

50% Ice concentration

75%

100%

Fig. 1.1: Color-coded ice concentration maps during maximum and minimum extents for 2014 in the (a, b) Northern Hemisphere and (c, d) Southern Hemisphere. The red contour represents averages using all data from 1978 to 2014.

can go as far south as 44° N. In the Southern Hemisphere, the sea ice cover surrounds the Antarctic continent and is generally symmetric in winter and on the average covers the Southern Ocean up to latitudes near 55° S. The sea ice in the Arctic is exposed to colder temperatures and is generally thicker than in the Antarctic. However, the ice in the Antarctic is more expansive, partly because no land boundary at the ice

1 Variability and trends of global sea ice cover and sea level

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edges limits its growth. The trends in ice cover in the two hemispheres are different: it is declining significantly in the Northern Hemisphere while it is expanding, albeit modestly, in the Southern Hemisphere. To some extent, this phenomenon is caused by differences in the geographical environment of the ice cover and the climate system of the two regions. It is, however, interesting to note that when the monthly values from the two hemispheres are added together for each month the sea ice cover shows a relatively uniform distribution with a slightly negative trend of about 2 % per decade.

1.2.1 Arctic Region The large-scale variability of the sea ice cover has been quantified using the terms ice extent and ice area. By definition, ice extent represents the integral sum of the area of all data elements in the map with ice concentration greater than 15 %. Ice area represents the sum of the products of the area of each data element and the ice concentration. In particular, ice extent represents the total area of the ice-covered region and provides the means to study how the fraction of the ocean with ice cover on it is changing. Ice area, on the other hand, provides the actual area covered by sea ice and the means to estimate the total ice volume assuming that the average thickness is known. The estimated ice volume can in turn be used for mass balance studies. A plot of monthly sea ice extents in the Northern Hemisphere from November 1978 to October 2014 is presented in Fig. 1.2 (a). The plot, which is an update of those reported previously [13, 27, 28] shows large seasonality of the sea ice cover with the winter extent as high as 16 × 106 km2 and the summer extent as low as 3.5 × 106 km2 . The plot also shows how the yearly seasonality of Arctic sea ice evolved from nearly uniform in the 1980s to a more variable one in the 1990s and to an even more variable one in the 2000s and 2010s. It is apparent that Arctic sea ice cover has become more seasonal in the last 8 years with sea ice cover at the end of the summer declining more rapidly than sea ice during the winter. To assess inter-annual variability and trends more quantitatively, monthly anomalies in the ice extents are presented in Fig. 1.2 (b). The anomalies were estimated by subtracting the climatological monthly averages (using averages of data from 1978 to 2014) from each monthly average. The plot shows high values in 1980 and 1996 and unusually low values in 2007 and 2012. The period from 1997 to 2006 also shows a moderate and steady decline but no unusual yearly changes. After 2006, the interannual variability became very strong in part because of inconsistent and relatively low values in the summer ice extent. In 2012, the monthly anomaly went down to a low value of almost −3 × 106 km2 . The results of a linear regression analysis also indicate an average decline of 3.8 percent per decade from 1978 to 2014. To gain better insight into the changes in sea ice cover, the maximum and minimum extents and areas of annual sea ice cover are presented in Fig. 1.3 (a) and (b), respectively. The 5-day running average of daily extents was used to estimate the max-

6 | Josefino C. Comiso Northern Hemisphere 20 Ice extent

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Fig. 1.2: Plots of (a) monthly averages and (b) monthly anomalies of sea ice extent in the Northern Hemisphere for the period from November 1978 to October 2014.

imum and minimum for each year. The running average is used to minimize the effect of daily variability in the extent due to temporal changes in wind direction and other factors. The maximum extent is shown to be declining at the rate of 1.9 % per decade while the minimum extent is declining at the rate of 11 % per decade. The ice area is declining slightly faster with the maximum and minimum trends being 2.2 and 11.8 % per decade, respectively. The actual numerical changes in area and extent and corresponding statistical errors are listed in the figures. The results show that although both cases show negative changes, there is a large difference in the change of ice extent during the winter/growth period and the change in ice extent during the summer/melt period. The change in winter is usually influenced mainly by surface temperature and wind circulation, while the change in summer is mainly controlled by the thickness of

1 Variability and trends of global sea ice cover and sea level |

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20 (a) Ice extent, northern hemisphere 15

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Year Fig. 1.3: Plots of yearly maximum (blue), minimum (red), mean (green) and seasonal ice (gold) of (a) sea ice extent and (b) sea ice area from 1979 to 2014 in the Northern Hemisphere.

the ice floes and their ability to survive the summer melt. A fraction of the thicker ice floes are advected out of the region mainly through Fram Strait, but the inter-annual changes are minor and it is not known how the ice fluxes have changed during the satellite period [29, 30]. The change in minimum ice cover represents a change in perennial ice cover, which is ice that survives the summer melt. Such ice has been known to be in existence for a long time in the Arctic and has been observed at least for 1450 years [5]. A continuation of a decline of 12 % per decade implies that summer ice could disappear during the 21st century. It may actually disappear sooner than expected, since actual data show more rapid decline than what modeling studies predict [7]. Studies of changes in the thickness of the ice have been undertaken using a combination of submarine, mooring and satellite data; there is a consensus that the average thickness is declining as well [31]. A quantitative assessment of the inter-annual variability of Arctic sea ice cover is presented in Fig. 1.3. The yearly averages (in green) do not show the fluctuations in the

8 | Josefino C. Comiso monthly plots, but the yearly trends in the ice extent and ice area are −3.71 ± 0.34 % and −4.32 ± 0.35 % per decade, respectively, which are similar to those derived using monthly anomalies (see Fig. 1.2 (b)). The values derived from monthly anomalies are usually used because of higher statistical accuracy. The differences between the maximum extents (in winter) and the minimum extent (in the previous summer) provides the mean to assess how the seasonal ice extent has been changing; these are shown (in yellow) in Fig. 1.3. The inter-annual changes are relatively small except in 2007 and 2012. The trends for seasonal ice extent and ice area for seasonal ice in the Arctic are −7.0 ± 1.3 % and −6.44 ± 1.2 % per decade, respectively. These values indicate that the pan-Arctic ice cover is becoming more and more seasonal. The trends in ice cover are not uniform spatially and can be very different for different regions and seasons as depicted in the color-coded image shown in Fig. 1.4. In winter and spring, the significant negative trends occur in the peripheral seas and especially in the Sea of Okhotsk, Baffin Bay, Barents Sea and Greenland Sea. It is interesting to note that in winter and spring the trends are positive in the Bering Sea. The trend in ice cover in the Central Arctic is almost zero because the ice cover is fully consolidated most of the time during this period. During the summer and autumn, the trends are strongly negative in the areas where the ice has retreated the most during the 1978 to 2014 period.

1.2.2 Antarctic Region The monthly variability in the extent of Antarctic sea ice cover as observed using passive microwave data is depicted in Fig. 1.5 (a). It is apparent that ice cover in the Southern Ocean is more extensive in the winter but less extensive in the summer than those in the Arctic region [14, 18]. The ice cover is thus primarily seasonal, and perennial ice is almost all of the second-year ice type because it tends to be advected out of the perennial ice regions every winter. In the first two decades, the ice seasons were generally very similar and typically would range in extent from 3.0 to 18.5 × 106 km2 . In more recent years, however, the range shifted slightly higher and was about 4 to 20 × 106 km2 in 2014. The monthly and inter-annual changes are more evident in the monthly anomaly plot shown in Fig. 1.5 (b). A large fluctuation in the monthly anomalies is apparent and indicative of large inter-annual changes for each month. The plot also shows the positive trend as described earlier. The result of regression analysis indicates a positive trend of 2.0±0.2 % per decade. In light of what has been observed in the Arctic sea ice cover, such a trend is unexpected and has been the subject of much research activity [33–37]. Some look at longer-term time series and indicate that the ice cover was actually more extensive in the 1940s and 1950s according to ship observations [38, 39]. The changes have been primarily regional with the most positive trends occurring in the Ross Sea; it is strongly negative in the Bellingshausen/Amundsen Seas.

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Northern hemisphere seasonal ice concentration trends 1979-2014

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Fig. 1.4: Color-coded trend maps of sea ice concentration for (a) winter, (b) spring, (c) summer and (d) autumn using satellite data from 1979 to 2014. Trends are indicated in yearly change in area (km) and in percentage change.

Plots of yearly maximum and minimum extents and areas are presented in Fig. 1.6. The yearly maximum is again determined using five daily running averages to minimize short-term effects. The maximum was relatively uniform in the first 20 years, but has been on the rise in recent years with the maximum exceeding 20 × 106 km2 for the first time in 2014 during the 1978 to 2014 period. This means that more sea ice was produced in the Southern Ocean in 2014 than in any of the previous years. Ice growth

10 | Josefino C. Comiso Southern hemisphere 20 Ice extent

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Year Fig. 1.5: Plots of (a) monthly sea ice extent and (b) monthly ice anomaly in the Southern Hemisphere for the period from November 1978 to October 2014.

is usually most intense and is facilitated by colder temperatures at the ice margins and by stronger winds off coastal polynyas where ice is produced continuously [40, 41]. A freshening of the water due to melting ice shelves or more iceberg calving would also increase the rate of sea water freezing [37]. The ozone hole has been postulated as the cuase for a deepening of the lows in the West Antarctic; this has been confirmed by numerical models [36]. This in turn causes stronger winds off the Ross Ice Shelf and higher ice production in the Ross Sea region. This is consistent with the unusually high rate of growth of sea ice in the Ross Sea in recent years. Higher ice production off the Ross Ice Shelf has been confirmed in a number of recent studies [35, 42]. Yearly minimum and maximum extents and ice areas as presented in Fig. 1.6 indicate that the trend is much higher for summer minimum (i.e. 4.2 % per decade) than the winter maximum (i.e. 1.4 % per decade). This suggests a cooling of the ocean (as

1 Variability and trends of global sea ice cover and sea level |

11

will be shown later) that allows more of the thicker ice types to survive the summer melt season. It may also mean less melting – but considering the large seasonality of ice in the region, the effect on the productivity of the region may be insignificant. The yearly overall averages are also presented and shown to have a slightly less expansive trend than the monthly anomalies shown in Fig. 1.6, in part because the 2014 yearly average was not included. The figure also shows a plot of the yearly seasonal ice extents (difference between the maximum and minimum extents) suggesting modest inter-annual changes in ice production. The trends for yearly and maximum extents are about the same at 1.4 % per decade; this indicates that the trend for yearly averages is heavily influenced by the long ice growth season that normally lasts for about 9 months. As mentioned earlier, it is important to note that the global sea ice extent using combined data shows a decline and a trend of −1.10±0.13 % per decade overall. 25 (a) Ice extent, southern hemisphere 20 15 10 Maximum trend: 26900 ± 6280 km2/yr Seasonal trend: 18500 ± 8360 km2/yr Yearly trend: 16800 ± 5450 km2/yr Minimum trend: 11000 ± 6180 km2/yr

Area (106 km2)

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Year Fig. 1.6: Plots of yearly maximum (blue), minimum (red), mean (green) and seasonal ice (gold) of (a) sea ice extent and (b) sea ice area from 1979 to 2014 in the Southern Hemisphere.

12 | Josefino C. Comiso

The trends of average ice concentration in the Southern Ocean for the different seasons are presented in Fig. 1.7. It is apparent that the trends are strongest at the marginal sea ice regions. It is also evident that there is an alternating positive and negative trend around the continent, but not consistently for the different seasons. The trends are indicative of the mode 2 or mode 3 patterns of sea level pressure that have been identified and studied previously [43, 44]. A pattern of a propagating wave called Antarctic Circumpolar Wave was identified in the 1980s and 1990s [41], but was not so evident in the late 2000s [31]. Instead, the wave became more stationary with more persistent positive trends in the Ross Sea region. Again, this is consistent with modeling studies indicating that the ozone hole had led to the deepening of the lows in the lower trophosphere in West Antarctica leading to strong winds off the Ross Ice Shelf [35]. Strong winds caused the formation of larger coastal polynyas in the region and the production of more ice.

1.3 Variability and trends in sea level One of the most serious impacts of global warming is the rise in global mean sea level (GMSL). Sea level rise is a big concern because a large fraction of the inhabitants of our planet lives in coastal areas. It has been estimated that an increase in sea level by a few meters would cause the displacement of several hundred million people, immeasurably large economic losses and a mass destruction of the environment and biodiversity. The environmental and ecological consequences are also expected to be profound [15]. In global mean sea level we refer to the height of the sea with respect to a benchmark (e.g. fixed reference such as a land feature) and averaged over long enough period to minimize if not eliminate the effects of big waves and tides. It also takes into account isostatic rebound as may be caused by the melt of ice sheets or glaciers over land and crustal movements. For a long time, the sea level has been monitored using tide gauges that were installed in cities and towns around the world. The gauges provide accurate and continuous readings of the sea level in areas where they are properly maintained. However, the measurements represent regional changes, and they don’t necessarily represent changes in GMSL. This problem was minimized if not eliminated with the advent of satellite radar and laser altimeters that provide ocean topography measurements at high precision globally. The rise in sea level is caused mainly by two factors: (a) a warming of the ocean; and (b) the introduction of liquid or solid water from land into the ocean. The melt of sea ice provides a negligible contribution to sea level rise because sea ice is already part of the ocean.

1 Variability and trends of global sea ice cover and sea level

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13

Southern hemisphere seasonal ice concentration trends 1979-2014

90°W

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% Ice concentration/year Fig. 1.7: Color-coded trend maps of sea ice concentration for (a) summer, (b) fall, (c) winter and (d) spring for the 1979 to 2014 period.

1.3.1 Contributions from warming oceans Liquid water is known to expand as its temperature increases on account of the enhanced kinetic energy of the water molecules; they thus require more volume for the same number of molecules. The ocean is a vast storage of heat and energy, and it is estimated that 90 % of the additional heat absorbed by the Earth in the last 50 years due to global warming has been stored in the ocean [45]. Getting estimates of the change in volume associated with global warming is not so straightforward, because of many complex ocean processes associated with atmospheric forcing, ocean dynamics and different physical characteristics of the ocean in different regions. The expansion of

14 | Josefino C. Comiso

the ocean depends on the quantity of heat absorbed and on water temperature, pressure and to a smaller extent, salinity. Greater expansion is expected to occur in warmer and saltier water under greater pressure. It has been estimated that for a simplified environment, sea water with a depth of 1 km expands by about 1 or 2 cm for every 0.1 °C increase in temperature [46]. If this estimate is correct, an increase of 4 °C for a doubling of the atmospheric CO2 could cause a significant sea level rise of 40 to 80 cm. However, it turned out that there are many complications that need to be accounted for in making such an estimate for ocean expansion in the global oceans [48]. Data on sea level rise as caused by thermal expansion has been relatively sparse until the 1980s when dedicated hydrographic measurements became available. Examples of repeated basin scale hydrographic measurements were those made for the North Atlantic [48, 49] and for the Southwest Pacific [50]. The results of these measurements as summarized [47] indicate that thermal expansion caused a sea level rise of about 1 mm per year. A major cause of uncertainties to estimates of the rate of sea level rise has been attributed to the occurrences of mesoscale eddies and large inter-annual variability in surface topography. 0.2

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Fig. 1.8: (a) Altimetry data from five groups (CU, NOAA, GSFC, AVISO, CSIRO) with mean of the five shown as bright blue line; and (b) yearly average GMSL reconstructed from tide gauges (1900–2010) by three different approaches as depicted in orange [61], blue [62], and green [63]. (With permission from IPCC and Cambridge University Press.)

Ocean thermal expansion has been cited as a major contributor to sea level rise in the 20th century; its effect is expected to continue in the 21st century [45, 51]. About half of the sea level rise over the last few decades has been attributed to the warming of the ocean [52]. Combined with contributions from other sources, data for sea level rise using modern techniques are presented in Fig. 1.8 (a), while the GMSL as reconstructed from tide gauges by different investigators are presented in Fig. 1.8 (b).

1 Variability and trends of global sea ice cover and sea level

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The data shows some yearly variability, but it is apparent that the rate of increase has been considerably higher in the last decade. The series of altimeter data starting with GEOSAT in 1985 and followed by the more advanced systems (the most prominent of which are the TOPEX and Jason series) are presented in Fig. 1.9. It is evident that there is a lot of noise in the data, especially with GEOSAT. There were also overlaps in measurements by the different sensors from 1992 to 2004; it is apparent that there are some inconsistencies. Such inconsistencies were resolved through the use of precise and well-documented reference frames (e.g. land). The long-term gauge data provided a rate of sea level rise of 1.32 mm per year while the altimeter data provided 2.7 mm per year for the period 1985 to 2004. Again, this is an indication of an accelerated sea level rise assuming that the two data sets provide consistent measurements of sea level. Data from TOPEX and Jason, which are very similar systems, shows sea level rise of 3.2 ± 0.4 mm per year for the period 1993 to 2009 [46]. 100

TTM Estimated global (81˚S to 81˚N) sea-level rise = 2.7 ± 0.4 mm/year

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Fig. 1.9: Global mean sea level as measured by different altimetry missions from 1985 to 2004 [64]. A precise and well-documented reference frame that is monitored for several years is essential to ensure the consistency of the measurements. (With permission from Wiley-Blackwell, LTD.)

1.3.2 Contributions from glaciers, ice sheets and others As indicated earlier, the melt of sea ice does not contribute to sea level rise because it is already part of the ocean. However, through ice-albedo feedback and other effects from the retreat of sea ice, especially in spring and summer, a general warming of the region occurs; this makes other components of the cryosphere more vulnerable. Sea level is affected by the transfer of mass from land to the ocean. The contributions from snow and permafrost are through river runoff; there has not been any indication of

16 | Josefino C. Comiso

significant increases in these contributions in recent years. The most important components of the cryosphere that could significantly, if not drastically, affect sea level are the glaciers and huge ice sheets in Greenland and Antarctica. The effect of climate change on glaciers is long-term and it is hard to make attribution of inter-annual changes in the volume because the changes may be the result of climate forcing from the previous century. But during the 20th century, the contribution of glacier melt to sea level rise has been estimated to be considerable and actually exceeded the contribution from ice sheets [51]. This is despite the fact that the combined volume of all glaciers represents only about 1 % of global ice volume. The big potential contributors to sea level rise are the ice sheet of Greenland which has a sea level equivalence of more than 7 meters and the Antarctic ice sheet which has a sea level equivalence of more than 60 meters. The impact of the loss of mass in glaciers is more than that associated with sea level increase. Glaciers help regulate the seasonal water cycle and provide fresh water to neighboring regions during the dry season. The retreat of glaciers also causes a destabilization of mountain slopes and sometimes leads to the formation of meltwater lakes that are unstable and can cause flooding. Attribution of the loss of glaciers to anthropogenic causes has been studied and found to be detectable only in recent years (1991 to 2010) at 69 ± 24 % of total loss when the anthropogenic signal becomes significant [53, 54]. The ability to assess the location of significant mass losses in Antarctica and Greenland has been made possible by new technologies. Many techniques have been adopted, including the use of radar altimetry, LIDAR altimetry and SAR interferometry; this results in estimates that are generally plausible but sometimes inconsistent [10]. Although data resolution is not as good as with other sensors, satellite data from the Gravity Recovery and Climate Experiment (GRACE) have allowed investigators to pin down the exact locations where the loss of mass is significant as illustrated in Fig. 1.10 (a) and (b). Thus in Antarctica, mass loss is most prominent in West Antarctica in the general location of the Pine Island glaciers; massive icebergs have been calving here in recent years. West Antarctica has been an object of intense research because the ice sheet is sitting on a bedrock that is primarily below sea level, and the average temperature during the summer is just a few degrees (≈ −6 °C) below freezing temperature. In Greenland, there is significant mass loss in an area of the southern and western region that occupies more than half of the total area. Quantitative assessments of the mass loss per year of the glaciers and those from Greenland and Antarctica are presented in Fig. 1.10 (c). It is apparent that at least up to the present, the contribution from glaciers is more than double those from either Greenland or Antarctica. The concern has been the stability of the ice sheets; it is currently not well established. In West Antarctica, there is concern that the intense calving in the Amundsen Sea region will continue if not accelerate and cause a considerably larger contribution to sea level rise. In Greenland, the area of the surface that experiences melt dur-

1 Variability and trends of global sea ice cover and sea level |

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Year Fig. 1.10: Distribution of mass loss expressed in cm of water per year for the period 2003 to 2012 as determined by the Gravity Recovery and Climate Experiment (GRACE) time-variable gravity for (a) Antarctica and (b) Greenland. The graph at the bottom provides the total mass loss of ice from glaciers and the Greenland and Antarctic ice sheets in giga-tons and sea level equivalent (mm) [52]. (With permission from IPCC and Cambridge University Press.)

ing the spring and summer has been increasing. The associated meltwater has been postulated to percolate to the bottom where it serves as a lubricant to the ice sheet and causes considerable changes in the dynamics of the system [53]. However, other scenarios are possible. For example, the meltwater could fill up the pores in the ice sheet and freeze; after that, it would be it unlikely that liquid water would percolate all the way to the bottom in subsequent melt periods. Fig. 1.11 shows yearly maps of areas that experienced melt during the year. It is apparent that the area of melt was most expansive in 2012 with almost the entire ice sheet having experienced some melt period. In 2002, the area of melt was also extensive, but not quite as bad as in 2012.

18 | Josefino C. Comiso Surface Melt in Greenland

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Fig. 1.11: (a) Coverage of annual maximum melt in Greenland using EOS/Terra using surface temperatures derived from moderate-resolution imaging spectroradiometer (MODIS) for the period 2000 to 2012; (b) actual area of melt with durations of one day, two days and greater than two days for each year from 2000 to 2012 [65]. (Reprinted with permission from Wiley-Blackwell/AGU.)

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1.4 Effects on physicochemical parameters The impacts of the observed trends in sea ice cover and sea level rise on the physicochemical characteristics in the regions are expected to be profound. The rapid retreat of sea ice cover in the summer will likely lead to a basically ice free Arctic Ocean with a totally different environment and ecosystem. We have already observed that warming in the Arctic is amplified; this makes the other components of the cryosphere vulnerable and will likely cause a rise in sea level. A most important impact would be the change in physical characteristics including the vertical structure and circulation of the Arctic Ocean. An important layer in the Arctic Ocean is the halocline, which has kept deep-ocean warm water from upwelling to the surface and melting the ice. Without the halocline, the physical characteristics of the Arctic Ocean will change, and habitats of the ocean will have to adjust accordingly. The impacts of sea level rise have direct and indirect components. Examples of the direct component are the displacement of hundreds of millions of inhabitants on the planet and the loss of important components of the climate system such as wetlands, mangroves and natural resources in the region. Among the indirect components are that a large fraction of species and organisms will likely disappear, and the biodiversity of the region will change considerably. Also, sea level rise will lead to the salinization of underground water near coastal areas; this is a critical source of fresh water and it is used in agricultural production. This leads to serious problems for hundreds of millions of people who depend on this source for their fresh water supply.

1.4.1 Large-scale changes in surface temperature Since 1900, the global surface temperature of the Earth as derived from meteorological stations and other data (i.e. GISS data) has been increasing at the rate of about 0.08 °C per decade [8]. For the period from 1981 to 2012, the same data shows a rate of increase of 0.17 °C per decade for the entire globe. This indicates an acceleration of warming (approximately two-fold) since 1981 despite a reported hiatus since 1998 that has been observed and hypothesized to be due to natural climate variability. The same data also shows a rate of increase of about 0.60 °C per decade for the Arctic region (> 64° N) while the satellite AVHRR data yields a rate of warming of estimated to be 0.69 °C per decade for the same region[8]. The results show that the rate of warming in the Arctic is more than three times the rate of warming globally. This is often referred to as the amplification of global warming in the Arctic [55]. The AVHRR trend value likely provides the more accurate assessment, because the data is more comprehensive and covers the entire Arctic region. Changes in water temperature have many implications for aquatic organisms at all trophic levels. Most organisms are sensitive to changes in water temperature, and most of them will not survive extremely high temperatures. Correlation analysis of the

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plankton concentration with surface temperature data has indicated that depending on the region, bloom patterns are enhanced in regions like the Bering Sea and Okhotsk Sea as the sea ice recedes; however, the plankton concentration starts to decline as the temperature reaches a certain threshold value. Increases in water temperatures to more than 29 °C have led to bleaching coral reefs in many parts of the world. It is also well known that the warming of Eastern Pacific Ocean during ENSO normally leads to the demise of millions of fish in the region. Tuna are usually found in waters that have temperatures of about 28 °C. Heat waves have occurred occasionally, and in Europe, a heat wave killed several thousands of people in August 2003 when 40 °C was reached and sustained for about a week. Also, changes in temperatures that are a few degrees above normal would usually shorten mating seasons and keep organisms from being able to reproduce. In the Arctic region, the color-coded trend maps of surface temperatures for the different seasons using satellite infrared data from 1981 to 2014 are presented in Fig. 1.12. Overall, the trends are dominantly positive, but it is surprising that there are regions in the Bering Sea, the western part of North America and Eastern Siberia where the trends are negative. The trend in the Bering Sea is consistent with the positive trend of sea ice extent in the region during the same period as independently measured. The trend is most positive where sea ice cover has been declining; this indicates that the temperature of surface water is significantly higher than that of ice-covered water, especially in the summer period. The trends in the eastern part of North America, which includes the Elsmere Islands where thousands of glaciers are located, are also strongly positive. This is also true in Greenland for most seasons. The strong positive trends in the Arctic Basis are consistent with the rapidly declining sea ice cover in the region. A similar color code trend maps of surface temperatures for the Antarctic region is presented in Fig. 1.13. The maps for each season obviously indicate less warming and more cooling than in the Arctic; this shows a consistency with the trends in the sea ice concentrations in the two regions. Vast areas of the ice-free part of the Southern Ocean shows some cooling, mainly adjacent to the sea ice cover. In Antarctica and sea ice-covered regions, a large fraction shows cooling trends except during the onset of melt temperatures in spring and when ice starts to melt. At this time, the sea ice cover has basically reached its maximum extent, and the warming trend may in part be associated with rapid breakup as could be caused by stronger winds. The more serious impact of such warming, however, would be more melt and an enhanced ice velocity for the Antarctic ice sheet and glaciers that would lead to increases in sea level.

1.4.2 Large-scale changes in plankton concentration and primary productivity The large-scale changes in the biology of the system can be best described in terms of satellite-observed plankton concentrations. In this regard, we combined data from

1 Variability and trends of global sea ice cover and sea level

60N 180E

90E

(a) Winter trend 1981-2014

(b) Spring trend 1982-2014

(c) Summer trend 1982-2014

(d) Autumn trend 1981-2013

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K/dec >2.05 2.0 1.9 1.8 1.7 1.6 1.5 1.4 1.3 1.2 1.1 1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0 –0.1 –0.2 –0.3 –0.4 –0.5 –0.6 –0.7 –0.8 –0.9 –1.0 –1.1 –1.2 –1.3 –1.4 –1.5 –1.6 –1.7 –1.8 –1.9 –2.0 2.05 2.0 1.9 1.8 1.7 1.6 1.5 1.4 1.3 1.2 1.1 1.0 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1 0.0 –0.1 –0.2 –0.3 –0.4 –0.5 –0.6 –0.7 –0.8 –0.9 –1.0 –1.1 –1.2 –1.3 –1.4 –1.5 –1.6 –1.7 –1.8 –1.9 –2.0 40. 30.

50°N 60°N 180°E

23

90°E

20.

10.

5. 3. 2.

Jan.

Mar.

May

1.

.5 .3 .2

.1

.05 .03 .02

Jul.

Sep.

Nov.

.01

Fig. 1.14: Monthly climatology of chlorophyll pigment concentration in the Northern Hemisphere every other month from January to November. The climatology is from data starting in 1998 to 2014 using SeaWiFS and MODIS data.

in productivity are apparent in all regions likely associated with changes in wind patterns that determine whether an upwelling of nutrients are going to occur or not. Changes in surface temperature could also be an issue, especially in the Baffin Bay/ Labrador Sea and North Atlantic regions. In the Southern Hemisphere, plankton distribution as depicted in Fig. 1.16 is more symmetric and appears to follow a circular pattern near the polar fronts and the Antarctic Circumpolar Current [5]. The bloom patterns are again most pronounced in spring and summer (see November and January). The most intense blooms occur off the shores of Argentina and are most likely associated with the availability of more nutrients and iron than the other regions. It is interesting that in winter, the regions immediately north of the sea ice cover have very low plankton concentrations. The lack of nutrients and iron in the regions during the period are probably the culprit. But ocean acidification, especially where it persists, is also a possibility. The influence of sea ice on the primary production in the Southern Ocean has been discussed in detail elsewhere [27, 57]. Plots of primary productivity in the South-

24 | Josefino C. Comiso

120 Eastern Arctic

Western Arctic Trend= 6.38 ± 4.63 (g C/m2/yr/dec)

100 80 60 Trend= 21.57 ± 6.38 (g C/m2/yr/dec) 40 (a) 300

(b) Bering Sea Trend= –10.95 ± 8.04 (g C/m2/yr/dec)

Sea of Okhotsk

250

200 Trend= 2.19 ± 7.50 (g C/m2/yr/dec) 150 (c)

Primary productivity (g C/m2/yr)

240 220

(d) Greenland Sea Trend= 9.14 ± 5.32 (g C/m2/yr/dec)

Borents Sea Trend= 28.28 ± 10.46 (g C/m2/yr/dec)

200 180 160 140 120 (e) 180 160

(f) Baffin Bay/Labrador Sea

Hudson Bay Trend= 13.37 ± 5.89 (g C/m2/yr/dec)

140 120 100 Trend= –1.04 ± 4.98 (g C/m2/yr/dec) 80 (g)

(h) North Atlantic

Average of Areas Trend= 13.37 ± 5.89 (g C/m2/yr/dec)

300 250 200 150 (i)

Trend= –14.33 ± 7.33 (g C/m2/yr/dec) 2000

2005

2010

(j)

2000

2005

2010

2015

Fig. 1.15: Plots of yearly primary productivity in different regions of the Northern Hemisphere as derived using Chlorophyll a concentration from SeaWiFS and MODIS data, AVHRR surface temperature and other parameters.

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45°E

30°S Jan.

135°E Mar. mg/m3 >30. 20.

10.

5. 3. 2.

1.

May

Jul. .5 .3 .2

.1

.05 .03 .02

.01

Sep.

Nov.

Fig. 1.16: Color-coded maps of monthly climatology of chlorophyll pigment concentration in the Northern Hemisphere for every other month from January to November. The climatology is from data from 1998 to 2014 using SeaWiFS and MODIS data.

ern Ocean are presented in Fig. 1.17. The data shows a negative trend for all sectors combined and also for each individual sector. This appears to be in contradiction to the positive trends in productivity in the region as reported previously for the period 1997 to 2008 [57]. However, a trend analysis of the data in Fig. 1.17 for the same period would also yield positive trends. This reveals a weakness of doing trend analysis on

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data with relatively short record length. The reason for the negative trends for the period from 1997 to 2014 is the consistently low values starting in 2011. The low values may be caused by the negative trends in the ocean’s surface temperature and the positive trends in the sea ice cover especially in the summer, but more research is needed to establish this. The enhanced values from 2001 to 2010 also need some explanation, but again – further studies are required. The ocean color data have been invaluable in many other marine science studies. For example, the data can be used to identify regions where harmful algal blooms (sometimes called red tides) occur. In this case, the signal is caused by a bloom of toxic red dinoflagellates. These are normally depicted as high values in plankton concentration, mainly in coastal waters. The data could therefore provide advanced and critical information about onset and magnitude of these blooms. Ocean color data are also invaluable in the assessment of the effects of ocean acidification as discussed in the next section. There are already some observations of areas that are highly depleted in plankton concentration and can be candidate areas for studying the effects of ocean acidification.

1.4.3 Changes in other physicochemical parameters The aforementioned changes in the global sea ice cover and the rise in sea level will undoubtedly cause pronounced shifts and reorganizations in global and regional ecosystems and biogeochemical cycles [59, 60]. Finding direct linkages of these changes to actual changes in physicochemical parameters (other than surface temperature and plankton concentration) is not trivial because of the lack of reliable numerical models that can be used for sensitivity studies and in situ data to validate the results of such studies.

1.4.3.1 pH A large fraction of atmospheric carbon dioxide is taken up by the ocean. After being absorbed, the carbon dioxide dissolves and forms carbonic acid, which breaks down quickly into bicarbonate (HCO−3 ) and a hydrogen ion (H+ ). The acidity of liquid water is usually quantified by its pH, which is a measure of its acid/base activity and defined as the negative common logarithm of the activity/concentration of hydrogen ions: pH = − log [H+ ] The pH of pure water at room temperature is seven and is regarded as the pH for natural waters that are not acidic. Fresh water with pH less than seven is considered acidic while those greater than seven represent base saturation or alkalinity. For ocean seawater, the pH level is slightly higher and is currently at 8.2. The introduction of more hydrogen ions would reduce the pH and make it more acidic.

1 Variability and trends of global sea ice cover and sea level

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Indian Ocean Trend= –5.22 ± 2.39 (g C/m2/yr/dec)

Weddell Sea Trend= –3.16 ± 2.23 (g C/m2/yr/dec)

30

25

20

15 (a)

45

(b) Ross Sea Trend= –6.46 ± 3.68 (g C/m2/yr/dec)

W. Pacific Ocean

Primary prodution (g C/m2/yr)

40 35 30 25 20 Trend= –3.16 ± 2.23 (g C/m2/yr/dec) (c)

60

(d) Average of Areas Trend= –5.81 ± 3.37 (g C/m2/yr/dec)

Bellingshausen, Amundsen Seas Trend= –9.01 ± 4.92 (g C/m2/yr/dec)

50 40 30 20 (e)

2000

2005

2010

(f)

2000

2005

2010

2015

Fig. 1.17: Plots of yearly primary productivity from 1998 to 2014 in different regions of the Southern Hemisphere as derived using Chlorophyll a concentrations from SeaWiFS and MODIS data, AVHRR surface temperature and other parameters.

Since carbon dioxide is taken up more rapidly in colder water, the regions that are most vulnerable are the polar regions. Thus, in the Arctic basin where sea ice is retreating, the presence of more cold open water areas that are being exposed due to ice cover decline means enhanced acidification in the region. The images in Fig. 1.14 and Fig. 1.16 show low plankton concentrations in parts of the ocean. Such low values may

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be in part a manifestation of acidification in the region, but some validation studies are needed to make sure that this is indeed the case.

1.4.3.2 Dissolved oxygen The availability of dissolved oxygen (DO) in the ocean is required by all forms of aquatic life. The concentration of DO is influenced primarily by biological activities and through photosynthesis in aquatic plants. Thus DO is relatively high during the daytime and is reduced during nighttime. Oxygen also tends to be more soluble in colder than warmer water and may be more readily available in the polar seas. The typical DO of unpolluted fresh water is about 10 mg/l. It has been observed that fish kills occur when the DO levels goes down to less than 2 mg/l. Again, a key source of dissolved oxygen is ocean plankton. Thus a depletion of plankton as may be caused by the loss of sea ice, ocean acidification or other factors would cause a reduction of dissolved oxygen.

1.4.3.3 Electrical conductivity and turbidity The electrical conductivity of water is mainly influenced by the amount of salt in the water. The more saline the water, the higher the electrical conductivity. Globally, the spatial variations of the ocean’s salinity is primarily controlled by precipitation, evaporation and sea ice cover. Precipitation causes the introduction of fresh water, which causes a reduction of salinity; evaporation does the opposite. In the polar oceans, the impact of sea ice on salinity depends on the season. During the growth period, the ocean in the underside of sea ice is usually cold and saline because of sea ice production. Sea ice can absorb only less than 30 % of sea water salt during formation; this causes the water on the underside of the ice to be relatively saline. This is especially the case where the rate of ice production is highest in autumn and winter as in coastal polynyas and ice edges. During the spring and summer, the opposite happens with the introduction of melt water that is less saline than the regular salt water. The positive trend in ice extent in the Antarctic means that the rate of ice production is getting higher and therefore more high-density saline water is being formed. On the other hand, the negative trend in the Arctic during the summer means a freshening of the surface water with the introduction of more meltwater. Turbidity or Total Suspended Solids (TSS) affect the transparency and light scattering of water and is usually composed of fine clay, silt particles, plankton, organic or inorganic compounds and microorganisms. The size of suspended particles ranges from 10 nm to 0.1 mm but are normally defined as those that do not pass through a 45 μm filter. In the Arctic, turbidity is usually affected mainly by river run-off, windinduced upwelling and changes in plankton concentration; in the Antarctic, winds and plankton concentration are the key factors. Changes in the acidity of the ocean could affect turbidity, since it changes the solubility of suspended matter.

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1.4.3.4 Nutrients There is large spatial variability in the distribution of nutrients in the oceans and in the polar regions; they may not be as readily available in some as in other regions. The most vulnerable areas are those that are farthest away from land. The loss of sea ice in the Arctic basin exposes more open water with surface layers consisting of meltwater that can be the platform for phytoplankton blooms. However, this would not materialize if nutrients were not available to support the blooms. Also, as the ice retreats to the deep ocean region, it is likely that these regions are depleted of nutrients. The reason for this is that when sea ice forms, the water on the underside becomes dense due to salinization and tends to go down to the deeper portion of the water. During the process, nutrients are entrained with the dense water into the deeper parts of the ocean where it may not be possible for them to resurface.

1.5 Discussion and conclusions The effect of global warming on sea ice cover and the global mean sea level is studied and the possible impacts to the physicochemical characteristics in the region is evaluated. Global warming is not uniformly distributed around the globe. The warming has been observed to be amplified by about three times and observed to be 0.69 °C per decade in the Arctic region (> 64° N) compared to 0.17 °C per decade globally; however, no amplification is observed in the Antarctic region. The impact of ice-albedo feedback is apparent in the Arctic where dramatic reductions in summer ice is observed and the extent of the perennial sea ice cover has been reduced to less than half its value over a span of three decades. The impact of the decline of sea ice in the Arctic is a general warming in the region that has caused the volume of glaciers to decline and the Greenland ice sheet to lose mass, which together with the warming ocean has caused an increase in the rate of sea level rise to about 3.2 mm per year from 1.32 as measured by the long-term gauge data. The trend in extent of the Antarctic sea ice cover continues to be positive, and currently at the rate of about 2 % per decade. The positive trend is consistent with a general cooling in the region except in some parts of the West Antarctic region where mass loss is significant. The impacts of the observed decline of sea ice cover in the Arctic can be profound and would affect the physical oceanography, the primary productivity, and the circulation patterns of the ocean and the atmosphere. The rate of sea level rise is accelerating but currently the rate is minor. However, the ice sheets are relatively unstable and could contribute to large changes in the global mean sea level in the future. The impacts of changes in sea ice and sea level on the physicochemical characteristics of the affected regions including the ecology, the environment and aquatic organisms at all trophic levels are also very negative and could get worst by the end of the 21st century.

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[63] Ray RD, Douglas BC. Experiments in reconstructing twentieth century sea levels. Prog Oceanogr 2011;91:496–515. [64] Kuo CY, Shum CK, Yi y, Braun A, Schroeter J and Wenzel M. Determination of 20th century global sea-level rise. Geophys Res Abstracts 2006;6:07741. [65] Hall DK, Comiso JC, DIGirolamo NE, Shuman CA, Box JE and Koenig LS. et al. Variability in surface temperature and melt extent of the Greenland ice sheet from MODIS. Geophys Res Lett 2013;40:doi:10.1002/grl.50240.

Begoña Espiña, Marta Prado, Stephanie Vial, Verónica C. Martins, José Rivas, and Paulo P. Freitas

2 New techniques in environment monitoring 2.1 Introduction Microalgae in marine, and cyanobacteria in both marine and freshwaters are able to produce harmful effects, including a broad range of phenomena referred to as “Harmful Algal Blooms” (HABs). In recent decades, scientists have observed an increase in the frequency, severity and geographic distribution of HABs worldwide. Recent research suggests that the impacts of climate change may promote the growth and dominance of harmful algal blooms through a variety of mechanisms including: – warmer water temperatures; – changes in salinity; – increases in atmospheric carbon dioxide concentrations; – changes in rainfall patterns; – increasing intensity of coastal upwelling; and – sea level rise. An adequate management of HABs requires the monitoring of microalgae. The global scale, the severity of some toxins derived by them and their relation to environmental phenomena such as eutrophication highlights the need of adequate monitoring tools for microalgae and cyanobacteria in order to control and prevent HABs. Such monitoring tools should allow the fast, accurate and specific detection of HABs species in monitoring programs; this offers important advantages such as understanding the presence of HAB species and their distribution and dispersion mechanisms, which also contributes to the prevention or mitigation of HABs’ harmful effects on human health, marine ecosystems and related economic activities [1]. There are several physicochemical parameters that have demonstrated an impact and relevance on HABs’ occurence and can be correlated with microalgae abundance in order to monitor and predict them. These parameters are usually measured in water quality control programs and include the use of methods (sensors, in most cases) to measure conductivity, pH, dissolved oxygen, salinity, temperature, depth, turbidity, ammonium and nitrate. In this chapter we will not focus on those parameters, but rather on environmental monitoring tools whose main objective is to provide information about the biological composition of waters and toxin profiles. Our main goal will be to describe recent advances in new technologies for analytical systems and biosensing (in lab bench analysis and in situ), including nanotechnology and molecular biologybased methods as well as their application to HABs monitoring and control.

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Environmental monitoring covers a large range of fields related to air, soil and water analysis, and it can take many forms and have many levels of technical sophistication. This chapter will cover the technological progress in biosensing that has led to improved detection capabilities in the sensitivity, accuracy and multiplexability of biosensors applied to waterborne toxins, in particular toxins produced by phytoplankton (e.g. dinoflagellates and algae). Particular attention will be devoted to the nanomaterials investigated as enhancing strategies applied to biosensors for waterborne toxins. In particular, gold and magnetic nanoparticles adapted to immunosensors will be reviewed as immobilization supports, signal amplifiers and signal probes. Additionally, novel molecular strategies based on genetic tools for DNA/RNA amplification and detection techniques will also be reviewed.

2.2 In situ harmful algal bloom monitoring The increase in the frequency of HABs’ occurrence makes their prediction and early detection an even more important concern in environmental monitoring. However, early detection is difficult and response time increases when monitoring relies only on in situ sample collection and further laboratory bench analysis. Taking this into account, new automated systems for HAB monitoring are necessary. It is important to clarify that currently HAB monitoring is based on the detection and/or quantification of significant increases in the phytoplankton biomass that can include, or not include, a toxinogenic HAB. In this sense, harmfulness of the exponential growth of phytoplankton and/or cyanobacteria is not always based on the production of toxin, since many species do not or – at least were not – reported to produce biotoxins, but their massive growth can imply other noxious effects on the ecosystem such as oxygen deployment, decrease in light penetration depth, nutrients depletion, etc. Nevertheless, the exponential growth of toxinogenic species gives a new relevance to the HABs that should be taken into account. Even so, very few automated monitoring systems can discriminate toxic HABs from the ones that are not toxic. Taking this into account, we will devote this section to new trends, advances and currently used automated monitoring systems for HABs while trying to focus, when possible, on technologies that could discriminate between toxinogenic and non-toxinogenic ones.

2.2.1 Optical remote sensing Remotely sensed data collection and interpretation is the front line in monitoring and forecasting algal blooms. The use of Earth Observation (EO) remote sensing data for the study of HABs has been the object of many advances and studies such as in the

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ABDMAP (Algal Bloom Detection, Monitoring And Prediction) project, a Concerted Action funded by the European Commission and carried out from April 1997 to March 1999 [2]. Several institutions in oceanic countries develop, implement and operate services using satellite EO data for monitoring HABs and other water quality parameters. To do so, the Nansen Environmental and Remote Sensing Center (NERSC) in Bergen (Norway) as well as the Monitoring and Event Response for Harmful Algal Blooms (MERHAB) Research Program from the National Oceanic and Atmospheric Administration (NOAA) from the USA use satellite imagery data collected from the European Space Agency’s MERIS and other EO sensors to forecast the HABs’ occurrence from numerical ocean models based on optical signals derived from diverse parameters such as chlorophyll a, yellow substance and sea surface temperature (SST). The MERIS instrument has been particularly effective in estimating biomass mainly due to its 300 m resolution, two-day repeat orbit and sufficient spectral bands. In mid-2015, Sentinel-3, part of a series of Sentinel satellites from Copernicus program, is scheduled for launching; it will include an Ocean Land Color Instrument (OLCI) that is based on the heritage MERIS instrument. The OLCI operates across 21 wavelength bands from ultraviolet to near-infrared and uses optimized cameras directed to reduce the effects of sun glint. However, despite the EO’s usefulness, there is no evident relationship between the concentration or the biological composition of an algal bloom on the one hand and, on the other hand, the observed reflectance spectrum from the water in which the bloom occurs. The observed reflectance spectrum depends, for one thing, on the pigments present in the bloom. There are three main groups of photosynthetic pigments: chlorophylls, carotenoids and phycobiliproteins (phycobilins). Common in all photosynthetic organisms that produce oxygen, chlorophyll-a is often used to calculate overall phytoplankton abundance, although it is highly variable depending on physiological conditions, degree of light adaptation and state of degradation [3–5]. There are other oceanic physicochemical parameters used to monitor and forecast microalgae growth by optical remote sensing that are even more imprecise in terms of determining the biological composition or harmfulness of the episode. However, they can help in early warning and tracking of HABs, above all, when combined with other technologies. This is the case of SST tracking by infrared emission; in oligotrophic and mixed waters, sun energy can penetrate as much as 30 meters in depth, implying a high temperature mitigation rate. However, when there are high concentrations of phytoplankton in surface waters, relatively little solar energy can penetrate the water column, meaning that most of it is absorbed as heat in the upper few meters. Additionally, SST allows for inference in the availability of some necessary nutrients for phytoplankton. Some of them, such as nitrite, are only supplied in normal conditions from the waters below the thermocline, mainly in upwelling areas. So, SST would act as a tracer of nitrite concentration [2]. It is also very important to keep in mind that

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availability or scarcity of certain nutrients may favor or disfavor the growth of certain phytoplankton species (toxinogenic or not). Taking this into account, a deep knowledge of favorable eco-physiological conditions for harmful species growth is essential. As reviewed, optical remote sensing systems constitute highly valuable tools to help and support the forecast and monitoring of the occurrence of HABs; nevertheless, currently these tools are still not absolute solutions as they have some limitations that can only be overcome by in situ sampling and analysis.

2.2.2 Automated monitoring Early warning systems are needed to prevent and minimize the consequences of HABs’ occurrence. Regular monitoring usually implies collecting samples at specific places and time points, and posterior lab bench analysis; e.g. optical microscopy for phytoplankton species identification and analytical chemistry procedures for toxin profiles in the case of contaminated seafood or freshwater. Those are time-consuming activities and mean high costs in terms of necessary instrumental and specialized staff. Due to this, frequency of sampling and processing, and, consequently, data delivery time lapse are not at the desirable standard. Ideally, early warning systems should be continuously working devices distributed in proper sampling points, able to collect and analyze representative samples of plankton and deliver specific data about their biological composition and toxinogenicity on a regular basis. In fact, there are already devices that are manufactured and available with approaches not so far from this challenging goal. McLane Research Laboratories, Inc. (Massachusetts, USA) produces automated technologies for in situ water sampling coupled with analysis. One of them, the Environmental Sample Processor (ESP; see Fig. 2.1), developed by the Monterey Bay Aquarium Research Institute, provides in situ collection and analysis of water samples from the subsurface ocean [6] (http://www.mclanelabs.com/master_page/producttype/samplers/environmental-sample-processor). The instrument is an electromechanical, fluidics system designed to collect discrete water samples, concentrate microorganisms or particles and automate the application of molecular probes in order to identify microorganisms and their gene products. Generated data are then available for remote retrieval and analysis in near real-time. The system is a modular design consisting of a core sample processor (the ESP), analytical modules and sampling modules. The core ESP provides the primary interface between the environment and a set of DNA and antibody-based sample processing technologies that are applied onboard the instrument in real-time. In addition, the ESP can be used to archive samples for a variety of analyses after the instrument is returned to a laboratory. The system provides expandability to allow installation and control of secondary analytical modules for parallel processing of collected samples. This system has been used to track HABs by analyzing water samples for the presence of toxic diatom species of

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the genus Pseudo-nitzschia, the dinoflagellate Alexandrium catenella, and the raphidophyte Heterosigma akashiwo. Additionally, an Enzyme-Linked ImmunoSorbent Assay (ELISA) test for domoic acid (DA) detection and quantification was included. This represents the first example of a complete automatic environmental sampler that can monitor the HABs’ producer along with its toxic product: Pseudo-nitzschia and domoic acid. [7]

Fig. 2.1: Environmental Sample Processor (ESP). Printed with permission from McLane Research Laboratories

Sample processing modules are able to perform quantitative Polymerase Chain Reaction (qPCR) to amplify DNA and/or detection by Sandwich Hybridization Assay (SHA) of the microorganisms present in the sample as well. This device has a version that can operate at a depth of 4000 meters, the deep-sea environmental sample processor (D-ESP). This is probably not so relevant for monitoring HABs because most of the toxin-producing phytoplankton is in surface waters. However, it has demonstrated its usefulness in surveying aerobic methanotrophs [8]. ESPs, even though they are promissory, present several drawbacks and issues that should be solved before being implemented in a systematic way as monitoring tools. ESP devices are huge and very heavy. This implies that deployment and mooring are very complicated and expensive procedures. Additionally, these systems are quite ex-

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pensive (more than $ 200 000) and present serious limitations concerning the number of samples that can be analyzed per deployment (typically up to 4 HABs consisting of several species and toxins). Additionally, the range of toxins and their producers that can be analyzed should be increased. So far, ESPs can only be used to detect DA and saxitoxin (STX) and their producers. Alternately, the Imaging FlowCytobot (IFCB) is an in situ, automated, submersible and imaging flow cytometer that generates images of particles in-flow taken from the aquatic environment down to a depth of 40 meters. This system would allow one to image all the planktonic cells present in seawater. The IFCB uses a combination of flow cytometry and video technology to capture high-resolution images of suspended particles. Laser-induced fluorescence and light scattering from individual particles are measured and used to trigger targeted image acquisition; the optical and image data are then transmitted to shore in real-time. Collected images during continuous monitoring can be processed externally with automated image classification software. Images can be classified to the genus or even species level with demonstrated accuracy comparable to that of human experts. IFCB generates high-resolution (1 μm) images of suspended particles in the size range < 10 to 100 μm (such as diatoms and dinoflagellates). The instrument continuously samples at a rate of 15 ml of seawater per hour. Depending on the target population, the IFCB can generate on the order of 10 000 high-resolution images per hour. However, the IFCB has some issues as well, mainly related to high power consumption, cost ($ 125 000) and sampling: only 15 ml/hour seems not to be enough to discriminate HAB species presence if there is a low abundance. Robotic instruments such as the ESP and IFCB show considerable promise for use with a range of HAB species. Both are commercially available as 2nd -generation instruments with 3rd -generation designs underway. However, both need strong testing and support from the HAB research and management community if they are to continue to develop further as products and as HAB research and monitoring tools. Correlating oceanic physicochemical parameters with changes in the planktonic population could help not only to forecast HAB events but also to understand the actual relevance of climate change on them. Automated sampling systems allowing spatio-temporal records has been increasingly useful in this regard. The Sir Alister Hardy Foundation for Ocean Science (SAHFOS), located in Plymouth, UK, is responsible for running one of the longest marine biological monitoring programs in the world, a survey on plankton dynamics with spatio-temporal records on a monthly basis in the North Atlantic Ocean and North Sea for over 60 years. The Continuous Plankton Recorder (CPR) is towed at a depth of approximately 10 meters. Water passes through the CPR and plankton are filtered onto a slow-moving band of silk (270 micrometer mesh size) and covered by a second silk. The silks and plankton are then spooled into a storage tank containing formalin. On return to the laboratory, the silk is removed from the mechanism and divided into samples representing 10 nautical miles (19 km) of tow [9].

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Nowadays, there are several institutions devoted to implementing Continuous Plankton Recorder Surveys (CPRS) all over the world. The Global Alliance of Continuous Plankton Recorder Surveys (GACS) is an international scientific organization that was established in September 2011 which aims to join efforts in this field to produce a global database [10] (http://www.globalcpr.org/). Data obtained by CPR allowed for very interesting studies correlating changes in oceanic physico-chemistry parameters with changes in the biological composition of the plankton in a time scale suitable to understanding the actual contribution of climate change to the higher impact of HABs’ occurrence observed in recent decades [11]. Hinder et al. generated cyanobacteria and dinoflagellates’ occurrence time series for over 50 years by analyzing almost 100 000 samples from the CPR survey [12]. Their results of over 12 cyanobacteria and nine dinoflagellates taxa showed a general decrease in dinoflagellates’ abundance and a clear increase in cyanobacteria. For the dinoflagellate taxa, abundance tended to decrease with increasing temperature and/or wind; the reverse pattern tended to be found for diatoms [12]. In this regard, results from this study are consistent with the last Climate Change Report from the Intergovernmental Panel on Climate Change (IPCC). This compilation of the most recent results about climate change indicates that global average warming of ocean waters from 1971 to 2010 is 0.11 °C per decade in the upper 75 m. Furthermore, wind stress increased during recent decades. It is very interesting that warming is more prominent in the Northern Hemisphere, especially the North Atlantic. This area was the focus for the study from Hinder et al. [13]. Of particular interest for HABs are the data obtained related to Prorocentrum spp. and Pseudo-nitzschia, the main producers of marine biotoxins responsible for Diarrhoeic Shellfish Poisoning (DSP), okadaic acid and dinophysistoxins, and Paralytic Shellfish Poisoning (PSP): DA. It is very interesting that an increase in the appearance and abundance of Pseudo-nitzschia was observed; however, Prorocentrum spp. became rare in the last decade. In the case of Pseudo-nitzschia, a clear correlation can be established between their relative abundance and the increase and spread of the HABs that they produce. However, the scarcity of Prorocentrum spp. contrasts with the incidence of HABs produced by those dinoflagellates. Related to that, it is important to clarify that the CPR does not sample very closely (< 10 km) to the coast and hence will not reveal locally specific patterns of plankton change (for example those caused by local eutrophication and local changes in circulation due to coastal development), but instead provides a view of broader-scale regional changes. This means a major drawback for HABs’ tracking because most of their impact is recorded in coastal seawaters. Thus, a possible explanation for the Prorocentrum spp. decrease in abundance within the CPR survey could be a general change in distribution of this taxa to waters closer to the coast where the impact of human-related factors such as hypertrophycation, ballast water release, etc. are dominant.

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As reviewed, recent advances allow automated acquisition of highly relevant data related to HAB occurrence and abundance. However, automated systems still present some drawbacks and limitations that still make hand sampling and lab bench analysis necessary: – Poor development of automated systems for biotoxin detection and quantification: there are still very limited strategies for detection and identification of the toxins on automated systems together with limited sample processing (cell disruption and toxin extraction); – Limitations for sample collection in the relevant localizations: some of the most harmful microalgae species are not planktonic. For instance, Ostreopsis spp. are epiphytic microalgae able to produce palytoxins, one of the most potent no-protein toxic compounds. Those dinoflagellates cannot be monitored by automated systems, currently mainly devoted to planktonic samples.

2.2.3 HABs sampling based on absorption As reviewed above, there are currently diverse sampling systems allowing the characterization of planktonic biological composition of water and, to a very limited extent, their toxin content. However, the collection and analysis of truly representative volumes from big masses of water is very complicated. Additionally, it is necessary to develop monitoring tools that allow for the correlation of toxin concentration in water with toxin concentration in seafood when the toxin content is far below the maximum levels in order to prevent big economic loss in aquaculture due to closures. In this context it is very important to keep in mind that mollusks for human consumption are able to filtrate up to 40 l/day, meaning that they are very effective and high-yield concentrators. Furthermore, collection of spatial-temporal comprehensive data is necessary in order to be able to predict the appearance of new outbreaks. In the last decade, new absorption techniques have been developed and implemented for in situ passively concentrated contaminants present in water for posterior analysis in the laboratory. Polar organic chemical integrative samplers (POCISs), consisting in a sequestration medium enclosed within hydrophilic microporous polyethersulfone membranes have been successfully used to sample polar organic chemicals in water [14, 15]. Other kinds of semi-permeable membrane devices have been used for monitoring polycyclic aromatic hydrocarbons [16]. For waterborne biotoxins, passive Solid-Phase Adsorption Toxin Tracking (SPATT) systems have demonstrated their usefulness in concentrating lipophilic toxins (okadaic acid, pectenotoxins, azaspiracids, dinophysistoxins and yessotoxins) to a level able to predict HABs affecting mussels days or weeks in advance [17]. The SPATT bags consist of batches of previously hydrated adsorbent resins inside a polyester mesh that is sewn with polyester thread. Bags are deployed in a frame attached to a weighted line at previously determined places and depths [18]. The nature, properties and behavior

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of the solid-phase adsorbent are the major issues in the development of SPATT devices. Several polymeric absorption resins have been used, but the best results in terms of adsorption and desorption rates were obtained with resins based on a matrix of styrene-divinylbenzene; DIAION1 HP-20, SEPABEADS SP850 and SEPABEADS SP825L, and more recently, Amberlite XAD761, which is made of a formophenolic matrix with methylol functionality, for more polar toxins such as DA [17]. The average pore sizes of these resins are 260, 38 and 90–125 Å, respectively. HP-20 always accumulated more OA than other resins and did not reach equilibrium within 72 h, probably due to its larger pore size. A recent study showed that HP-20 can hold a maximum of 1639 mg/g of OA [19]. The same kind of resins have been used as the base of active sampling systems consisting of a submersible pump connected to a pre-concentrating device based on filtering by a mesh for planktonic collection to get rid of the bigger particles and cells. This processed sample is filtered again and divided to pass through different cartridges containing the HP-20 resin. This system allows for continuous pumping for even one week in a row; only filter cleaning is required. Using these devices during HABs, McCarthy et al. recently reported to have isolated, concentrated and purified as much as 13 mg of OA, 29 mg of DTX2, 20 mg of PTX2 and 6 mg of PTX2-SA during a seven-day deployment [20, 21]. Fewer studies have been conducted in the use of SPATT for freshwater toxin monitoring so far, probably due to the difficulties in finding a good adsorbent material for such small and hydrophilic compounds. However, powder-activated carbon G-60, Strata-X (a hydrophilic–lipophilic polymeric resin), AG 50W-X4 (strong cation exchange) and Amberlite IRP-64 (weak cation exchange) showed promise. Polycarbonate membranes with Oasis1 HLB from Waters (hydrophilic-lipophilic polymeric resin) seem to present good results for microcystin-LR (MC-LR) and microcystin-RR (MC-RR) [17]. However, Zhao et al. reported good adsorption rates of microcystins (MCs) for aromatic resins HP20 and SP700 in cyanobacterial cultures [22]. In contrast, sampling by adsorption can be used to simplify the collection and pre-treatment of samples for posterior analysis. This is the case of Solid Phase MicroExtraction (SPME), a sample preparation technology that uses a fiber coated with a liquid (polymer), a solid (sorbent) or a combination of both. The fiber coating removes the compounds from the sample by absorption in the case of liquid coatings or adsorption in the case of solid coatings. The SPME fiber is then inserted directly into the chromatograph for desorption and analysis. This technology has already been successfully used to pre-concentrate and/or extract hydrophilic toxins (saxitoxin, anatoxin-a and microcystin) from the environment for posterior analysis [23–27].

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2.3 Liquid chromatography and mass spectrometry When doing bibliographic research, it becomes clear that today the preferred analytical chemistry methods for the detection and quantification of water and food contaminants are based on chromatography coupled, or not, to mass spectrometry. As a consequence of their quick development and the interest in avoiding the use of Mouse BioAssay (MBA), High Performance Liquid Chromatography (HPLC) or Liquid Chromatography coupled to tandem Mass Spectrometry (LC-MS/MS) are, since very recently, the official detection methods in the European Union for most of the toxins regulated in legislation for both marine and freshwaters [28]. These methods are currently used in regular monitoring; however, they have been of vital importance for basic research in detecting and elucidating the structure of new toxins in natural samples as well. Thus new methods based on these technologies are developed every year and great efforts and resources are invested on research involving them. High performance liquid chromatography is basically a highly improved form of column chromatography. In this case, the solvent is forced to pass through a solid column under high pressures of up to 400 atmospheres or up to 1000 in the case of Ultra Performance Liquid Chromatography (UPLC). This makes it much faster. It also allows one to use a very much smaller particle size for the column packing material which gives a much greater surface area for interaction between the stationary phase (column) and the molecules flowing through it. This allows for a much better separation of the mixture’s components. The time that a molecule takes to get out of the column is known as the retention time, and it is a characteristic of that molecule in certain conditions. Using an Ultra Violet (UV) detector such as Diode Array Detector (DAD) or fluorescence lamp and detector after the column, the presence of one compound can be assigned to its retention time, separated from the mixture and quantified. In contrast, mass spectrometry (MS) determines the mass of a molecule by measuring the mass-to-charge ratio (m/z) of its ion. Ions are generated by inducing either the loss or gain of a charge from a neutral species. Once formed, ions are electrostatically directed into a mass analyzer where they are separated according to m/z and finally detected. The result of molecular ionization, ion separation and ion detection is a spectrum that can provide molecular mass and even structural information. Coupling LC to MS detection, identification and quantification of all the components of a mixture becomes virtually possible provided that proper molecular standards were available. Additionally, the sample preparation procedures are, in general, much simpler and faster than for LC. Mostly recent advances in MS technology allowed the recent development of new sensitive and reliable methods for toxins by using triple-quadrupole mass analyzers (QqQ), ion trap mass analyzers (IT), time-of-flight mass analyzers (ToF) and hybrid linear ion trap-Fourier transform mass spectrometers [29]. QqQ is currently the most popular type of detector because its sensitivity and selectivity in selected reaction monitoring mode (SRM) fulfills the necessities for de-

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tection and quantification for a high range of compounds. However, High Resolution Mass Spectrometry (HRMS), i.e. ToF and Linear Ion Trap Orbitrap (LIT-Orbitrap), provides promising results; operating in full-scan mode, making no pre-selection of SRM transitions, virtually all compounds present in a sample can be determined simultaneously. This allows one to carry out suspect and even non-target based screenings while avoiding the use of reference standards and saving data about the complete composition of one sample that can be further re-analyzed for the presence of new compounds of interest. Additionally, hybrid instruments offer the possibility of datadependent MS/MS acquisition; this means that an MS/MS is triggered only when a compound from a target ion list is detected in full scan. The use of QToF is a little bit restricted in quantification methods by the fact that the sensitivity and dynamic ranges are below the ones from QqQ. However, LIT-Orbitrap instruments offer the advantages of both technologies. Their main limitation so far is that these kinds of equipment are still very expensive, though less so than a few years ago. Due to recent developments in HRMS, a shift is taking place from parent compound analysis, mainly using multiresidue methods, to the identification of metabolites and transformation products [30–32]. In Europe, the Commission Regulation No 15/2011 amending regulation (EC) No 2074/2005 establishes the EU-Reference Laboratory LC-MS/MS method as the reference method for the detection of lipophilic toxins for both official control and own-checks. This includes the regulated groups of lipophilic toxins: okadaic acid, yessotoxins, pectenotoxins and azaspiracids. However, ELISA, HPLC-UV, LC-MS and UPLC-MS still coexist for the official control of Amnesic Shellfish Poisoning (ASP) toxins [28, 33, 34] while MBA and prechromatographic oxidation and LC-Fluorescence Detection (FD) are still used for PSP toxins [35, 36]. In regards to freshwater toxins (cyanotoxins), the trends are the same as for marine toxins. There are a high variety of developed and efficient detection and quantification methods using LC/HPLC/UPLC coupled to FD/UV. However, the higher sensitivity and specificity in identification is obtained when using LC coupled to MS, especially in tandem [30, 37]. This becomes of particular relevance for freshwater toxins as the limit of detection and quantification has to be lower to avoid pre-concentration steps. For instance, for MC-LR, the only globally regulated freshwater toxin so far, the guidance/regulatory limit is 1 μg/l [38]. Even despite it being less sensitive, accurate and selective than current LC-MS techniques, some groups of toxins are still officially analyzed by those methods as reviewed above. This is sometimes due to the lack of inter-laboratory validation for the most recent methods and, above all, because of the high cost of this equipment and for specialized operators.

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2.4 Biosensors for HABs monitoring Biosensors applied to waterborne toxins, in particular toxins produced by phytoplankton (e.g. dinoflagellates and algae) [29, 39] represent one of the scarce topics in the environmental biosensors area that might be economically favorable. The presence of biotoxins in water and seafood not only poses a health risk [40] to consumers but may also adversely impact the aquaculture industry economy by lowering consumer confidence in harvested shellfish products. With the increase of both toxin-producing organisms and seafood demand, the problem will only become more serious. This makes accurate and rapid detection of marine toxin-producing species and their respective toxins, both in water and seafood, increasingly urgent. Limits have been set for these toxins, and these vary from country to country along with the accepted methods for detection and quantification. Regulatory entities from the different countries stipulate both maximal allowable levels of toxins in domestic, exported and imported seafood products as well as the appropriate and legally acceptable methods to test for toxins in seafood products [41]. This is particularly relevant for the major classes of shellfish-poisoning according to their form of actuation, namely, Neurotoxic Shellfish Poisoning (NSP), Diarrhetic Shellfish Poisoning (DSP), PSP, Amnesic Shellfish Poisoning (ASP) and Ciguatera Fish Poisoning (CFP) [42]. New toxins are continuously being added to this list, most recently azaspiracid, and the “fast acting toxins” (gymnodimine, the spirolides and pinnatoxins). Currently available methods for monitoring marine toxins include the goldstandard, US Food and Drug Administration (FDA) approved, MBA, along with analytical techniques (e.g. chromatography), immunoassays and functional methods such as cytotoxicity assays, receptor-binding studies and enzyme inhibition studies [43]. Regarding the MBA, there are both ethical and technical concerns associated to testing procedures, as well as concerns about inadequate detection limits for the prevention of long-term chronic effects. Also, the MBA is unable to identify the exact toxin or even class of toxins involved in a poisoning event on a molecular level. The latter methods present distinct advantages and disadvantages in terms of specificity, ability of toxin identification, time of analysis, specialized staff and knowledge required, technical limitations and cost of operation depending on the type of toxin under analysis; but in general, all of them suffer from a lack of uniformity in their inter-laboratorial outcomes [44, 45]. Quite a few recent publications, including full journal special issues, cover this topic for the most relevant biotoxins under surveillance and for the respective detection methods that are either in development and/or already being implemented [46]. Compared to traditional detection methods, biosensors promise outstanding performance with higher selectivity, sensitivity and stability. They are faster and less expensive, but above all their in situ field analysis, miniaturization, portability and automation are the most appealing features [47, 48]. Usually the compact size of these detection systems enables in situ real-time analysis, avoids sample transportation and

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degradation, and provides almost immediate interactive information about the tested samples. They thus allow for a prompt decision and adoption of most adequate preventive and corrective measures before commercialization and dissemination of contaminated products and an occurrence of toxin outbreaks. This section is devoted to advances and new trends in biosensing tools development applied to toxins detection related to HABs’ monitoring. The original idea of a biosensor, where part of a living organism – the bioreceptor – generates a signal to monitor or measure the presence of a substance, was disclosed approximately six decades ago [49]. However, the most illustrative example of this concept dates back to the early 20th century, when coal miners used canaries to detect toxic gasses inside of mines. Here the bird represents the bioreceptor. Since then, the research on biosensors for environmental monitoring applications has been designed to address a critical growing need for in situ, cost-effective and disposable monitoring devices that are capable of quickly and accurately detecting an increasing number of contaminants. However, specific requirements must be met for each field of application. In the case of environmental biosensors, the US Environmental Protection Agency (EPA) made a rough estimate of the ideal range of specifications for a biosensor to succeed in this field of application. Those specifications include: cost of analysis between $ 1–15; portability by one person, with no need for external power; assay time within one to 60 minutes; limited personnel training needed, the device can be operated after 1–2 hours of training period; operation format continuous and in situ; minimal sample preparation for groundwater, soil extract, blood and urine; sensitivity of parts per million to parts per billion; dynamic range of at least two orders of magnitude; and a defined specificity for bioreceptors where enzymes, receptors and nucleic acids are specific to one or more groups of related compounds and antibodies specific to a single compound or closely related group of compounds. The International Union of Pure and Applied Chemistry (IUPAC) defines a biosensor as a self-contained integrated device that is capable of providing specific quantitative or semi-quantitative analytical information using a biological recognition element (the bioreceptor), which is retained in direct spatial contact with a transduction element (the sensor). The already numerous possible combinations of the various bioreceptors (e.g. DNA, antibodies, enzymes, organelles, bacteria, virus, cells, etc.) [50] with different transducers (e.g. electrochemical, potentiometric or amperometric, optical, thermometric, piezoelectric, etc.) to generate biosensors, continue to increase along with the discovery of new physical sensors. Examples of quite recent entries for the list of transducer types are magnetic sensors [51, 52]. The progress of biosensors is mainly, but not exclusively, dependent on advances and innovations of the previous two elements, the transducer and bioreceptor, which are directly related to the fast-growing areas of nanotechnology and biotechnology, respectively.

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On the one hand, biotechnology allows one to tailor the bioreceptor molecules or even create new ones (e.g. synthetic receptors and biomimetic ligands) with enhanced properties such as selectivity, affinity and stability, enabling better performance at the bio-recognition level. As an example, molecularly imprinted polymers (MIPS) are completely artificial binders based on synthetic monomers [53]. Briefly, a template molecule (analyte) is mixed with one or several functional monomers and a porogen (solvent) and a polymerization catalyst or starter. In most cases, a high concentration of a cross-linker (multivalent monomer) is also added. After the polymerization is finished (e.g. by heating), the material is ground and extracted thoroughly to remove most of the template molecules. The remaining molecular cavities can be used for selective binding of the respective analyte. Robustness is the great feature of these bioligands, which can stand very harsh conditions (e.g. acidic and alkaline media, organic solvents) without loss of functionality, as easily happens to protein-based bioreceptors. On the other hand, continuous developments in the nanotechnology area are promoting the rapid miniaturization and improved performance of sensors, but also serving the development of other nanomaterials, such as gold nanoparticles (AuNPs), carbon nanotubes (CNTs), magnetic nanoparticles (MNPs), quantum dots (QDs) and dendrimers, among others, that are able to improve biosensory capabilities [54]. Nanotechnology advances applied to biosensing will be reviewed in depth in the next section, which is especially devoted to the use of nanoparticles. A third relevant element is the support material for the immobilization of the bioreceptor molecules where characteristics such as ease of fabrication, special optico-physico properties, tailored morphological structure and versatile surface chemistry are desirable. Amongst materials being used as bio-sensing substrates, nano-porous silicon has gained popularity in recent years [55]. Also the application of biomimetic nanochannels in biosensors, as single nanochannels or nanochannel arrays, seem to bring new advantages for biosensor development and applications in environmental analysis [56]. All this in combination with improved integrated microfluidics [57, 58] and electronic platforms [59] is allowing specialists to develop a new era of smarter and exceptional devices that are able to assist in monitoring the quality and stability of our environment. The micro-total analysis system (μ-TAS) or “lab-on-a-chip” concept has been existent for approximately one decade; here several laboratory processes are integrated that include all sample preparation steps, incubation, detection and data analysis on a single device. However, even with all the progresses made, the identification of different analogues from a same group of toxins remains an unattained challenge even for the actual biosensing methods [60]. Also important to notice is that, despite the wide range of biosensors developed so far or under development with potential for applications in toxicity screening, relatively few of these have evolved into commercial devices [61].

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This chapter will cover the different types of biosensors and the major technological progress already attained or being pursued in the endeavor of waterborne toxin monitoring.

2.4.1 Optical biosensors Typical optical schemes are based on simple absorption spectroscopy (from the UV to the deep infrared), Raman and conventional fluorescence spectroscopy and imaging, but also on more sophisticated methods such as surface plasmon resonance (SPR), evanescent wave or fiber optic spectroscopy. Like other sensory principles, optical biosensors take advantage of rapid advances in nanotechnology and instrumentation. Current studies that focus on optically based transduction methods aim to achieve a more robust, easy-to-use, portable and inexpensive analytical system.

2.4.1.1 Surface Plasmon Resonance One of the methods that has most of the required biosensing capabilities is SPR. This technology became popular in the 1990s when biosensors were commercialized by the Swedish company Biacore® . The phenomenon of surface plasmon resonance allows for the study of molecular interactions in real-time and without the need of molecular labeling. The association of SPR technology with nanostructured materials is even increasing its capabilities. The conjugation of the analyte with nanoparticles and/or the modification of the SPR gold layer with multifunctional nanomaterials is being used to enhance the refractive index changes. For example, the use of magnetic nanoparticles results in a signal amplification effect and higher sensitivity [62]. Also, AuNPs functionalized with different polymers, such as dendrimers and borin acid, have proved to amplify the SPR response and lower the detection limits [63–65]. SPR-based biosensors are presently being investigated as food and environmental analytical tools [66], in particular when applied to the complex problem of marine toxins [67, 68]. Researchers at Queen’s University in Northern Ireland had pioneered the use of SPR biosensors to detect marine biotoxins. Early research was performed developing a biosensor assay for DA [69] and later for more complex toxins, such as DSP [70, 71] and PSP [72, 73]. Those toxins, DA and OA, were detected in the range of ng/g of the analyzed sample, while PSP toxins were detected in concentrations of two to 50 ng/ml. In general, these assays use antibodies as bioreceptors (immunoassays) to perform inhibition or competitive strategies. The binding of the antibody to the standard toxin on the sensor surface is inhibited by the presence of the target toxin in the sample solution, and the signal is inversely proportional to the sample concentration. This same competitive mode is the basis of a novel automated online optical biosensing system (AOBS) developed for the rapid detection of MC-LR. In this system, water samples

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containing the MC-LR were pre-mixed with a certain concentration of fluorescencelabeled anti-MC-LR monoclonal antibody, which binds to MC-LR with high specificity and results in a detection limit of 0.09 μg/l [74]. Dendrimers are also being explored to improve the performance of SPR-based biosensors. When used as an intermediate layer, dendrimers overcome the mass transport-limiting effects in the ligand immobilization; this increases its efficiency. Furthermore, dendrimer-based affinity sensor matrices were found to exhibit greater stability and regeneration of bioreceptor layer [75, 76]. Lately, multiplex SPR biosensor systems for the detection of multiple toxins have suddenly begun to appear. A first attempt to develop a model system compatible with a Biacore prototype multiplex analyzer utilized a microfluidic immobilization device for the covalent attachment of up to 16 different bioreceptors for four toxin groups in an array on a single surface [77]. Later, a modification of this system for the simultaneous detection of PSP toxins, okadaic acid (and analogues) and DA was developed. The prototype’s detection limits based on Inhibitory Concentration 20 (IC20) values for PSP, okadaic acid and DA toxins were improved to the sub ng/ml level [78].

2.4.1.2 Lateral Flow Immunoassay On the other side, Lateral Flow Immunoassay (LFIA) is one of the simplest optical devices for detecting the presence of a target analyte in a sample, often by visual inspection of the results, without the need for dedicated reading equipment. LFIA combines the sensitivity and selectivity of immunoassays, such as ELISA, with simplicity and portability; it is among the rapid methods for first-level screening of food and environmental contaminants. It has recently attracted scientific and industrial interest because of the attractive property of enabling very rapid, one-step, in situ analysis [79]. Here, as was reported for SPR technology, the combination of the detection system with nano-engineered particles makes it benefit from their unique optical properties and highly improve the assay sensitivity. The conjugation of the target analyte with a labeling particle is used to overcome the limitations of traditional LFIA; nanoparticlebased systems have achieved notable progress in signal amplification strategies and improved sensing performance. Recently a new concept of LFIA, the so-called high performance lateral flow immunoassay (HP-LFIA), was introduced and validated to detect the ASP toxin, DA [80]. The test provides a qualitative result to indicate the absence or presence of toxins in extracts of shellfish tissues but at concentrations that are relevant to regulatory limits. An improved version of the test uses an electronic reader to remove the subjective nature of the generated results and employs simple extraction and test procedures, which require minimal equipment and materials. The analysis time is fast; it presents results in about 15 minutes after sample preparation and is suitable to be performed in remote locations.

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2.4.1.3 Chemiluminescence ChemiLuminescence ImmunoAssay (CLIA), first reported in the late 1970s [81, 82], simultaneously combines the high sensitivity of chemiluminescence analysis and the good specificity of immunoassays. Chemiluminescence reactions use labeling molecules (luminol, dioxetane, derivatives of acridine and imidazol) that lead to an emission of light under the action of various chemical and enzymatic compounds. A CLEIA was developed for the determination of MC-LR in water samples. Under optimum conditions, the calibration curve obtained for MC-LR had a Limit Of Detection (LOD) of 0.032 ng/ml and the quantitative detection range was 0.062–0.65 ng/ml [83]. However, immunoassays in general, and ELISA in particular, are difficult to handle and rather time-consuming. Therefore, different strategies have been pursued to attempt to overcome this limitation. One of these, flow injection systems are based on a very simple instrumental set up of easy automation, has attracted a great deal of attention. In this context, a portable chemiluminescence multichannel immunosensor (CLMADAG) based on a capillary ELISA technique in combination with a miniaturized fluidics system was developed [84]. The device (despite a not highly sensitive LOD above 0.2 μg/ml) could analyze a single sample of MC-LR in just 13 minutes. Besides being fast, it was a very reproducible and had an test format that was easy to perform.

2.4.2 Electrochemical biosensors Electrochemical biosensors are currently among the most popular of the various types of biosensors. This is mainly due to their extraordinary limit of detection. An up-todate review with an in-depth description of the electrochemical biosensors for the analysis of toxins, in particular marine toxins and cyanobacterial toxins that incorporate nano-biotechnological concepts, can be found in Campas et al. [85]. The most common improvements in these systems are associated with the use of carbon nanomaterials (e.g. grapheme, carbon nanotubes, nanohorns and nanospheres), which have been used as components in electrochemical biosensors for over a decade mostly due to their electronic properties.

2.4.3 Mass biosensors 2.4.3.1 Quartz crystal microbalance Quartz Crystal Microbalance (QCM) is a label-free piezoelectric mass sensing technique. In simple terms, it gives a response that characterizes the binding event between the analyte to be detected and a sensing layer, which is immobilized on the

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surface of the QCM transducer. The resonant QCM frequency depends on the mass attached to the quartz crystal surface. Standard QCM is ideal for detecting analytes of high molecular weights, however upon different optimization strategies and/or nanoparticle enhancement, it has also proven to work for smaller analytes, such as cyanobacterial toxins [86]. Thus, various amplification methods have been developed; in particular, AuNPs have attracted attention due to their unique physical and chemical properties, including easily controllable size distribution, long-term stability and biocompatibility with immunospecies.

2.4.3.2 Cantilever biosensors Cantilever sensors have received considerable attention over the last decade because of their potential as a highly sensitive sensor platform for high throughput and multiplexed detection. One particular type of cantilever sensor is the piezoelectric-excited millimetersized cantilever (PEMC), which is a macrocantilever composed of two layers: a PZT (lead zirconate titanate) and a nonpiezoelectric layer (glass) of a few millimeters in length. The PZT layer is attached to the cantilever’s base and acts both as an actuating and as a sensing element, while the non-piezoelectric layer provides a surface for bioreceptor immobilization. PEMC sensor is used in resonance mode. When the target analyte binds to a bioreceptor immobilized on the cantilever sensor, the effective mass of the cantilever increases and alters the sensor resonance frequency. Therefore, measuring the resonance frequency provides information not only about the presence of the antigen but also its concentration. Reports have been published on the detection of MC-LR in the dynamic range of 1 pg/ml to 100 ng/ml [87].

2.4.4 Magnetic-based biosensors 2.4.4.1 Nuclear magnetic resonance (NMR) A stable and sensitive toxin residues immunosensor based on the relaxation of magnetic nanoparticles was developed. The method was performed in one reaction and offered sensitive, fast detection of target toxin residues in water. The target analyte, MC-LR competed with the antigens on the surface of the magnetic nanoparticles and then influenced the formation of particle aggregates. Accordingly, the magnetic relaxation time of the magnetic nanoparticles was changed under the effect of the target analyte. The calibration curve was deduced at different concentrations of the target analyte. The LOD of MC-LR was 0.6 ng/g and the detection range was 1–18 ng/g. Another important feature of the developed method was easy operation: only two steps were needed to mix the magnetic nanoparticle solution with the sample solution and read the results through the instrument. Therefore, the developed method may be a useful tool for toxin residues sensing and may find widespread applications [88].

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2.4.4.2 Magnetoresistive biosensors This type of biosensor makes use of highly sensitive micro- or nano-structured magnetoresistive (MR) sensors associated to super-paramagnetic nanoparticles as reporter systems [51, 89]. MR sensors are made of nanometric multilayers stack, intercalating magnetic with non-magnetic thin-film materials. The electrical resistance of a MR sensor varies with an applied external magnetic field due to the scattering of “spin up” and “spin down” electrons at the interfaces of the multilayer sensor stack. This variation is linear in a relatively short magnetic field range (i.e. 50 Oe range), conferring quantitative sensing capacity. Spin valves (SV) are now the most frequently used type of MR sensors applied to the detection of biological events. SV are a particular type of MR sensor composed of two ferromagnetic layers, one with a fixed magnetization and the other free to rotate with an external magnetic field, separated by a non-magnetic metal spacer. The angle between the magnetization directions dictates the resistance of the sensor. When they are in a parallel configuration, the resistance is minimal; an anti-parallel configuration produces maximal resistance. In a standard MR biochip-based bioassay, a recognition bioligand immobilized over the sensor is used to interrogate an unknown sample potentially containing a target molecule of interest (e.g. toxin), labeled with a MNP. Whenever there is recognition between the target and its bioligand, a biomolecular event occurs. After washing, the recognized targets stay over the sensor, while the unbound molecules are washed out. Applying an external magnetic field, the magnetic labels attached to the bound molecules will create a fringe field further detected by the MR sensor [52]. In addition to fast response times, this technology is extremely sensitive, down to the level of single molecule detection. The magnetic nanolabels are discretely identified by the MR sensors thus avoiding the “yes or no” qualitative type of answer. Another important asset is the multiplexability, which detects the presence of several analytes in the same sample [90, 91]. Nevertheless, fully integrated and compact devices are still missing to reach the market and be implemented as standard methods. In this sense, the department of Environment Monitoring, Security and Food Quality Control from the International Iberian Nanotechnology Laboratory-INL [92] (www.inl.int) is addressing the development of nanotechnology-based detection technology for multiple cyanotoxins using the MR platform previously developed at INESC-MN group [89].

2.5 Advances in nanotechnology for HAB detection As discussed above, environmental monitoring needs rapid, accurate, easy-use, low-cost and miniaturized analytical tools for the detection of marine toxins. Nanotechnology [93–95] offers interesting features to address these criteria and lead to a breakthrough in the development of sensitive and reliable biosensors. Nanotech-

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nology implies the manipulation, creation and investigation of systems in a dimension range of 1 to 100 nm in which nanomaterial plays an important role. Inorganic nanoparticles (NPs) [96] present unique and remarkable physicochemical properties (optical, magnetic, electronic . . . ) that differ from those of the corresponding bulk materials and make them suitable for the development of biosensors. Moreover, their ease of synthesis, high surface area and high ability to be functionalized with a variety of biomolecules and small molecules [97] allow for a great increase in the amount of immobilized biomolecules and chemical reactions, thus improving the sensitivity and speed of detection responses. In the context of sensing, NPs are categorized mainly as a platform to generate a physical signal, as well as a signal amplifier [95]. In recent years, a multitude of highly sensitive detection and imaging methods for medical, food and environmental applications using NPs has emerged. Integrating NPs within biosensors during their fabrication processes have showed distinct advantages by combining low cost and ease-of-use with high sensitivity and selectivity. Therefore, the use of NPs for the detection of pollutants such as heavy metals [98, 99] toxins have been studied recently in the field of environmental monitoring [50, 100]. Whereas a large number of NP-based detection assays have been developed, their use for the detection of marine toxins is still in the early stages. This section illustrates the recent approaches using gold and magnetic NPs, which are the most common nanomaterials investigated for the detection of waterborne toxins in immunoassay systems. Moreover, we have categorized three strategies: NPs as immobilization supports; NPs as signal amplifiers; and NPs as signal probes.

2.5.1 Nanoparticles 2.5.1.1 Gold nanoparticles One of the most widely used NPs in the nano-biotechnology area is AuNPs since they provide a bio-compatible micro-environment for biomolecules [101]. Moreover, in the development of immunoassays, AuNPs have the advantages of being easily functionalized with the molecular capturing agents or the molecular target that favors the biorecognition interactions between the two components. Exploiting the distinct optical properties of AuNPs is a good alternative in terms of sensitivity, easy use and low cost for the development of immunoassays, diagnostics and biosensors of various analytes [100, 102]. These physicochemical properties are derviced from the localized surface plasmon resonance (LSPR), a collective oscillation of the conduction electrons that, for spheres, typically occurs in the visible to near-UV region spectrum. Excitation of these plasmon leads to a strong band around 520 nm. A small change in the size, shape, local environment, surface nature, inter-particle distance and degree of aggregation of AuNPs leads to tunable changes in their optical properties. Anisotropic gold nanorods (AuNRs) as an alternative to spherical NPs have also been studied and showed interesting features that potentially make them more

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versatile as a platform for biodetection [103, 104]. The AuNRs present two surface plasmon resonances: longitudinal and transverse [105]. The Longitudinal Plasmon Band (LPB) maximum is, typically, in the near-IR region; its position is highly sensitive to the local environment making it very attractive for biosensing [106]. Moreover, using AuNRs with different aspect ratios may provide a novel optical multiplex biosensor platform, and has broad potential applications in immunoassays. It is known that well-dispersed solutions of AuNPs display a red color, while blue color appears when they aggregate. This optical phenomenon is associated with the shift of the surface plasmon band of AuNPs as a consequence of interacting surface plasmon and aggregate scattering properties. Therefore, colorimetric sensors exploiting the optical properties of AuNPs have been widely developed for the detection of toxins, heavy metals [98] and other environmental pollutants in water, soil and other environmental samples [50] and showed distinct advantages such as high sensitivity and selectivity.

2.5.1.2 Iron oxide nanoparticles Magnetic NPs exhibit remarkable physical properties and have been exploited in a variety of bio-applications including hyperthermia, magnetic resonance imaging, contrast agents, tissue repair, immunoassay, drug/gene delivery and cell separation [107]. Due to their biocompatibility, low toxicity, simple preparation, low cost and superparamagnetic properties, magnetite Fe3 O4 and maghemite (𝛾-Fe2 O3 ) have received considerable attention for their use in biosensory applications. Functionalized with biomolecules such as proteins and antibodies, they have been served as a separation platform under external magnetic fields with the aim of concentrating and purifying the analytes. Their large surface area allows them to be used as a solid support for biomolecules in order to enhance the specific capture of the targeted biomolecules. Therefore, the use of NPs as biomolecule carriers in analytical chemistry is promising [108].

2.5.2 Analytical nano-applications 2.5.2.1 Nanoparticles as immobilization support In the aim of improving the detection limit of immunosensors, the use of NPs as antibody immobilization supports offers interesting advantages. Because of their high surface area, NPs provide a way to increase the amount of immobilized antibodies and improve stability and reusability of the system. Therefore, the antibody-antigen recognition event is favored and faster; it also enhances the sensitivity of the biosensors. One of the strategies consists of modifying the working electrode with the NPs which leads to a large increase in antibody loading on the sensor. Among the wide varieties of biosensors, mainly electrochemical immunosensors have recently reported

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the detection of waterborne toxins while taking advantage of this strategy. Metallic NPs such as gold or silver NPs have been integrated into electrochemical sensors since they both facilitate the electron transfer reactions with electro-active species and simplify the process of immobilization of antibodies on the electrode surface. Loyprasert et al. [109] have developed a capacitive immunosensor for the detection of MC-LR. Silver NPs were fixed on the electrode surface using a thio-urea monolayer and were functionalized with an MC-LR antibody. Then, the antibody-antigen binding was directly indicated by capacitance measurement. The immunosensor showed a good sensitivity with a LOD of 7.0 pg/l and could be reused up to 43 times; this reduces the cost of detection. Likewise, L-cysteine AuNPs were deposited on the surface of gold electrodes for the immobilization of MC-LR antibody [110]. Upon addition of the MC-LR target analyte, the electrochemical impedance and the differential pulse voltammetry (DPV) were measured and linear detection was observed from 0.05 to 15 μg/l with a detection limit of 20 ng/l. The biosensor was validated with the analysis of MC-LR in spiked crude algae samples and showed 95.6–105 % of recovery; this makes it potentially suitable for field analysis. Recently, Lebogang et al. [111] have prepared a capacitive immunosensor by surface modification of gold electrodes with AuNPs deposited on polytyramine layers. The monocolonal antibody immobilized on the modified surface allowed the capture of the MC (≈ 80 congeners) via the Adda-moiety. The stable biosensor showed a linear detection in a range of 1 × 10−13 to 1 × 10−10 M of MC-LR with a LOD of 2.1 × 10−14 M with the ability to be reused up to 30 times. Due to their large surface area and unique super-paramagnetic properties, MNPs have also shown a considerable potential to be used as immobilization support in sensing. Moreover, they can be easily fixed and maintained on the working electrode when applying a magnetic field. Hayat et al. [112] used protein G magnetic beads to develop a label-free amperometric immunosensor for the detection of OA. The detection was carried out by DPV in a ferri/ferrocyanide reaction. An anti-OA monoclonal antibody was bound to the magnetic bead to form a complex. When the electrode was incubated in presence of OA, the specific affinity antigen-antibody led to the formation of layers on the electrode surface. This phenomenon generates an electron resistance between the redox couple. This system showed a LOD of 0.5 μg/l which was comparable to the results obtained with ELISA in terms of sensitivity. The experiments with mussels confirmed the applicability of the immunosensor for OA detection. Similarly, coreshell Fe3 O4 @Au magnetic NPs were immobilized onto the surface of the screen-printed working electrode using a permanent magnet and were functionalized with MC-LR antibody via adsorption (Zhang et al.) [113]. The as-prepared sensor was based on the direct competitive immunoassay format between the labeled agent of horseradish peroxidase (HPR)-conjugated MC-LR and the MC-LR (target analyte). The capture of MC-LR was measured by DPV. The peak current response decreased proportionally with the concentration of MC-LR in the range of 0.79–12.9 μg/l with a detection limit of 0.38 μg/l. The immunosensor, validated in the real water samples, provided fast response and good sensitivity.

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Tang et al. [114] have designed a magneto-controlled electrochemical competitivetype immunoassay of brevetoxin (BTX) in seafood. Guanine-assembled graphene nanoribbons (GGNRS) were used as molecular tags on a home-made magnetic carbon paste electrode. Previously, monoclonal anti-BTX antibody was immobilized on magnetic beads. In absence of BTX-2, the magnetic bead is captured by the bovine serum albumin BTX (BTX-2-BSA) attached to GGNRS. However, by increasing the amount of BTX-2, the amount of magnetic bead that interacts with GGNRS decreases, resulting in a decrease of electrochemical signals. This system exhibited a dynamic concentration range that spanned from 1.0 pg/ml to 10 ng/ml with a detection limit of 1.0 pg/ml BTX. The second strategy consists of using nanoparticles as a carrier platform for antibodies, exerting an amplification effect and improving the detection limits of the antigens. Garibo et al. [115] have used magnetic NPs as antibody immobilization supports and carriers and as signal amplification for the detection of OA using an SPR device. The detection was performed following a direct competitive immunoassay. The antibody conjugated to protein G-coated magnetic NPs competes with the target analyte (free OA) and OA fixed to the SPR device. Compared to a free antibody, the conjugates enhanced the SPR signal up to 11-fold, which made it possible to utilize the antibody economically. The use of conjugates in the assay provided 3-fold lower LOD μg/l (2.6 μg/l, equivalent to 12 μg of OA/kg mussel meat). The system was validated following the analysis of mussel samples obtained from Ebro Delta’s bays during a (DSP) event. The immunochemical SPR approach has been demonstrated to provide specific and real-time monitoring of OA with minimal sample and reagent consumption in short analysis times. Another group proposed a rapid sandwiched immunoassay for MC-LR with flow injection chemiluminescence detection and using PEI-modified magnetic beads (PEI-MBs) as capturers and HRP-functionalized silica nanoparticles as signal amplifiers. When using polyethyleneimine (PEI)-magnetic beads (MBs) as the carrier of anti-MC-LR, the chemiluminescence signal was greatly enhanced up to 9-fold compared to one using MBs without PEI modification. The signal was further amplified 13-fold when Si-antibody was used as the signal probe. Under optimal conditions, the immunoassay exhibited a wide quantitative range from 0.02 to 200 μg/l with a detection limit of 0.006 μg/l [116]. On the other hand, Tang et al. have developed an immunosorbent bio-barcode assay [117] for the detection of gonyautoxin 2/3 (GTX 2/3). Here, the free target antigen and the fixed antigen GTX 2/3-glucose oxidase (GTX 2/3-GOX) on ELISA plate competes to interact with dual-modified antibodies – double strand DNA AuNPs. Once GTX 2/3 is captured by the 13 nm AuNPs via antigen-antibody affinity, the mixture is separated from the solution and heated for 5 min at 100 °C to release the single strand DNA that are not attached to the particles. The oligonucleotide, used as a template for signal amplification for bio-recognition events, was combined with PCR technique for quantification. The final results showed a LOD of 0.74 μg/ml. The developed sensor is a rapid and high-throughput tool to detect marine toxins. Multi-detection of paralytic, diarrheic and amnesic shellfish toxins by an inhibition immunoassay us-

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ing microsphere-flow cytometry system was reported [118]. Three toxins in solution compete with their analogues immobilized on the surface of three different classes of fluorescent magnetic microspheres (luminex). Therefore, the antigen-antibody recognitions generated three different fluorescent signals. The inhibitory concentration 50 (IC50) obtained in the buffer was similar in single and multidetection: 5.6 ± 1.1 ng/ml for STX, 1.1±0.3 ng/ml of OA and 1.9±0.1 ng/ml for DA. The three immunoassays performed well with mussels and scallop mixtures displaying adequate dynamic ranges and recovery rates (amount 90 % for STX, 80 % for OA and 100 % for DA). The method provided an easy and rapid screening method capable of simultaneously detecting three regulated groups of marine toxins in the same sample.

2.5.2.2 Nanoparticles as a physical amplification signal In order to detect trace levels in real samples, amplification strategies are thus required to improve biosensors’ sensitivity. Integrated NPs in such devices allow for the enhancement of the generated signal and improve the detection limit. QCM biosensors are widely used in immunoassays because of their low cost, high portability and their ease of use. However, in order to enhance their sensitivity for the detection of trace biological targets, AuNPs were added to the system as a mass amplifier. The target analyte is captured both by the antibody immobilized on the electrode surface and by the antibody anchored to NPs. The AuNP’s mass significantly affects the vibrational frequency of the quartz crystal that enhances the detection signal. A good example is the highly sensitive, portable method developed for detecting MC-LR using AuNPs. The method includes an inexpensive sensor, composed of a QCM along with AuNPs, and a sandwich immunoassay for rapid and in situ detection of MC-LR. By using this method, a detection sensitivity of 0.1 ng/l of MC-LR can be reached; this meets the standard of World Health Organisation (WHO) requirements for drinking water (1 ng/ml of MC-LR) and is compatible with conventional techniques, such as HPLC and ELISA. The size of the gold nanoparticles clearly influenced the amplification efficiency of the MC-LR signal. It was found that 30 nm in diameter is the optimal particle size [119]. Han et al. [120] have reported a rapid, inexpensive, in situ and highly sensitive method for detecting MC-LR. A sandwiched QCM immunoassay sensor was developed and consisted of three steps: (1) immobilization of the antibodies on the sensor; (2) antigen-antibody recognition events; and (3) incorporation of 30 nm AuNPs. In absence of AuNPs, the system could detect no more than 1 ng/μl whereas the addition of AuNPs exhibited a LOD of 0.1 pg/μl. On the other hand, Xia et al. [121] have shown the enhanced sensitivity of QCM via double amplification for the detection of MC-LR. As a first amplification, a vesicle matrix functionalized with an optimized MC-LR antibody was deposited at the surface of the QCM and allowed for high coverage and effective immobilization of the antibody. The results showed that a detection limit of 100 ng/ml was achieved, but the LOD was lower than the WHO’s required value. How-

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ever, when conjugated anti-MC-LR/AuNPs were added as a secondary amplification system, higher sensitivity with an LOD of 1 ng/ml was achieved and confirmed the need to use particles to provide higher sensitivity devices. AuNPs have also been exploited to improve electrochemical immunosensors. For example, Tang et al. [122] have developed such a sensor for the fast screening of BTX B (BTX-2) in food samples. Two nanomaterials were used: AuNPs to improve the conductivity; and dendrimers to increase the surface coverage of the biomolecules on the electrode. The competitive type immunoassay format using HPR-labeled anti-BTX antibodies exhibited high sensitivity, acceptable reproducibility and selectivity. A low detection limit of 0.01 ng/ml and a wide dynamic working linear range of 0.03–8 ng/ml BTX-2 using AuNP-dendrimer as matrices were obtained. In this study, simple sample preparation steps beyond extraction, filtration and dilution were performed. This method provided a biocompatible immobilization and a promising immunosensory platform for analytes with small molecules in food safety analysis and detection. Zhang et al. [123] have used AuNPs to enhance the performance and sensitivity of an electrochemical immunoassay based on novel capillary electrophoresis (CE) for the detection of shellfish toxins. CE separation power was improved due to the change of apparent mobility provided by the interaction between analytes and AuNPs. Enhanced resolution and sensitivity were obtained by using bioconjugates featuring HRP-labeled antigen linked to AuNPs in a competitive immunoreaction format introduced into a capillary system. The LOD were in the range of 3.1–36.7 ng/l successfully applied to the simultaneous determination of four toxins [123]. Improving the biorecognition procedure by an additional separation step, the LOD was lowered to 0.1 ng/ml [124].Here they demonstrated that the use of AuNPs enhanced the sensitivity of the simultaneous detection of the four shellfish toxins in shellfish samples more than 70-fold.

2.5.2.3 Nanoparticles as a physical signal Many nanomaterials are considered an ideal signal-generating probe due to their intrinsic physicochemical properties. In the case of AuNPs-based biosensors, mainly electrochemical, colorimetric and fluorescent devices were investigated. A few examples for the detection of marine toxins are described here. Due to their electronic properties, AuNPs were integrated in a porous paper-working electrode for the development of an electrochemical biosensor where particles are exploited as the label to produce the signal. Consequently, Ge et al. have developed such a system for the detection of MC-LR [125]. The target analyte was captured in a sandwich between antibodies immobilized on the electrode and those attached via click chemistry to 60 nm magnetic silica NPs combined with HPR. The magnetic silica NPs serve as signal amplification. For verification, the specificity of the assay was evaluated using other MCs variant (MC-YR, MC-RR, MC-LW) and nodularin toxins. No specific current change was observed for MC variants and nodularin toxins, indicating that those toxins did not in-

60 | Begoña Espiña et al.

terfere in the MC-LR detection. Thus, this assay provided the potential to quantify the MC-LR levels in a range of 0.01–200 μg/ml with a LOD of 0.004 μg/ml. Based on the optical properties of gold NRs, Sun et al. [126] have developed an LSPR assay for the simultaneous and multi-detection of pefloxacin (PEF) and MC-LR in seafood. Here, two aspect ratio AuNRs with an LPB located at 695 nm and 863 nm were functionalized with a PEF-antibody and MC-LR antibody, respectively. In order to enhance the signal and allow rapid separation, biological magnetosomes that provide good dispersivity, biocompatibility and super-paramagnetic properties, were functionalized with antigen-ovalbumin (A-OVA). The addition of the magnetosomes in AuNRs solution allows them to aggregate, resulting in a significant red shift of LPB. However, in the presence of free antigens, the competition between A-OVA and the target analytes reduces the aggregation states. With an increase in target concentration, the LPB shift is less pronounced and gradually decreases. The difference in shift between the peak positions of LPB in the absence and presence of PEF and MC-LR was proportional to the targets’ concentrations and allowed for their quantification. The use of magnetosomes have shown a response amplification which is 2.5–5-fold with a good linear response over 1–20 ng/ml. Compared with these methods, the sensitivity of the proposed method is good; furthermore, the new method described in the present study has shorter detection times (the entire analysis take less than 3 h), and the developed method is simple to operate because of the use of magnetosomes. In addition to playing a role of direct optical probes, AuNPs have also been exploited as a fluorescence quencher to develop optical sensors for target detection. Based on an energy donor and acceptor pair of graphene oxide (GO) and AuNPs which lead the fluorescence quenching GO, Shi et al. [127] have developed a reliable and highly sensitive immunosensor for the detection of MC-LR. The sensor was constructed as follows: (1) deposition of a GO sheet to a positive slide in an array format; and (2) attachment of a specific antibody to the Adda group of MC onto the GO layer. In parallel, the MC target was immobilized to the AuNPs (MC-AuNPs) via interaction with single strand DNA. Upon the addition of MC-AuNPs to the sensor, the recognition event between the antibody and MC allows interaction between GO and AuNPs. Therefore, the distance between the two components is close enough to quench the photoluminescence of GO by fluorescence resonance energy transfer (FRET). By increasing the concentration of MC, i.e. increasing the number of particles on the GO surface, the fluorescence intensity decreased. This system showed LOD was 0.5 and 0.3 μg/l for MC-LR and MC-RR. Moreover, they performed the detection of shellfish poisons produced by cyanobacteria such as STX and neoSTX using the same process, and no fluorescence quenching was observed, which highlights the selectivity of their sensor. The development of a lateral flow immunochromatogtaphic (LFIC) test strip assay using AuNPs as optical molecular probes has been exploited as an analytical method for the detection of waterborne toxins. This technique, which is based on a competitive immunoassay, provides a fast, easy and on-site qualitative visual response (yes/no result) of the presence of a target analyte for a wide range of medical, food and envi-

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ronmental applications, and is well known as a home pregnancy test. The principle consists of the migration of the sample (in presence or absence of the target analyte) and specific monoclonal antibody-conjugated AuNPs (mAb-AuNPs) through capillary force along the membrane strip. In an absence of the target analyte, mAb-AuNPs bind to both a competitive antigen at the target in the test zone and to an antibody such as IgG protein in the control zone. The accumulation of red-colored AuNPs in the test and control zones provokes the appearance of two clear red lines in the strip. However, in the presence of the target analyte, the target is recognized via mAb-AuNPs and the number of particles that can interact with the competitive antigen decreases, resulting in a test line with a weaker intensity. Colloidal gold particles were used in an immune-chromatographic assay strip for the detection of MC, and when compared with the unmodified analogous assay, it presents the strong advantage of a direct detection of MCs at the test site without having to bring the samples back to the laboratory for test analysis [128]. Zhou et al. [129] developed a screening method for the rapid detection of tetrodotoxin (TTX) in puffer fish tissues using the LFIC method. 50 μl of sample are absorbed and the response is obtained with a visual detection limit of 40 ng/ml 10 minutes later. The principle is based on the competition between the antigen (TTX-BSA) that is immobilized to the pad and the target analyte (TTX) in solution with monoclonal antibody-conjugated AuNPs (mAb-AuNPs). TTX-BSA and goat anti-mouse IgG are separately stripped onto the control and test region. When no TTX are present, the colloidal mAb-AuNPs is recognized by both TTX-BSA (test line) and goat IgG (control line). Therefore, two distinguishable lines (test and control) appear on the strip due to the accumulation of AuNPs during the recognition events. However, in presence of the toxin, TTX interacts with mAb-AuNPs and thus the amount of particles captured by TTX-BSA decreases. Color density is reduced with increasing TTX concentrations. Recently, the same group [130] reported an LFIC test strip for the detection of total OA and dinophysistoxin (DTX-1) in shellfish products. The principle was the following: OA-BSA and goat anti-mouse IgG are separately deposited onto the strip as a test line and control line, respectively. When no toxins are present in the sample, the monoclonal antibodies, conjugated to 20 nm AuNPs (mAb-AuNPs), interact with both the OA-BSA and the anti-mouse IgG, leading to the appearance of two distinguishable redcolored lines (test line and control line). However, when the toxin is present, OA is captured by mAb-AuNPs. Therefore, with an increase in the sample’s OA concentration, the number of particles that can interact with OA-BSA decreases, resulting in a weaker test signal. A positive signal is observed when the test line is barely visible and only the control line is visible. The qualitative detection limit of the 150 μg kg−1 sample was close to the European Union regulatory limit (160 μg kg−1 ). This assay showed that 10 minutes (except sample preparation) were needed to detect the toxin and that numerous samples could be screened. Moreover, the results obtained were confirmed by HPLC-MS/MS. This makes the LFIC method convenient and efficient as a preliminary screening tool for food safety.

62 | Begoña Espiña et al.

Liu et al. developed an LFIC for detection of OA with LOD of 5 ng/ml [131]. The analysis data found were in agreement with those obtained by ELISA. The developed mAb was immobilized to the AuNPs. In this study, the negative results were confirmed by the presence of two red lines: one in the test zone (test line) when mAb-AuNPs is captured by OA-KLH, and the other one in the control zone when mAb-AuNPs interact with rabbit antimouse IgG. With the absence of OA in the sample, the maximum of particles react with OA-KLH, the test red line is therefore intense. They presented better results than those obtained by Lu et al. [130], where LOD was 30–50 ng/ml (see Fig. 2.2 and Fig. 2.3). In their study, optimum conditions have demonstrated that the use of 40 nm particles gave better visibility compared to 20 nm, probably due to their optimal size and their lower steric hindrance. The IC50 value of ELISA was 0.077 ng/ml, and the gold nanoparticle immunostrip had a visual detection limit of 5 ng/ml. The OA strip is a semi-quantitative, rapid and visual assay indication to examine OA in shellfish samples within 10 minutes. This strip will be conceivably applicable for one-site detection of OA in the seafood industry. Based on the yes/no results of the LFIC, Jellet Rapid testing Ltd. has commercialized rapid, simple and point-of-care tests to identify ASP, DSP and PSP in phytoplankton and shellfish. The ASP and DSP kits show a sensitivity in shellfish of 0.075–0.1 μg/g OA that is in good agreement with the European Union regulatory limit (0.16 μg/g). The PSP shows a sensitivity in shellfish of 20–70 μg/100 g and a sensitivity of 0.02–0.1 μg/ml. Therefore, the LFIC allows the rapid screening of a wide range of samples and can be used as a first test to reduce the number of samples that need to be tested; this reduces the cost and time spent. Because of their super-paramagnetic properties, magnetic NPs have been used directly for the detection of MC-LR. Ma et al. [88] have reported the development of an immunosensor based on nuclear magnetic resonance. Here, the MC-LR target analyte competes with its analog which is attached to magnetic NPs (MC-MPs); this influences the aggregation between particles. In absence of the target, the particles tend to aggregate when a secondary antibody is added to the mixture and thus induces a magnetic relaxation time (T2). However, in the presence of free MC-LR, the target competed with the antigen attached to MPs, which inhibited the aggregation of magnetic aggregates, causing a change in the T2 values and magnetic resonance image. This method allowed sensitive and fast detection of the target toxin in one reaction, and the LOD (LOD) of MC-LR was 0.6 ng/g in a detection range of 1–18 ng/g. This method was sensitive and simple, including only two steps: mixing the magnetic NPs in the sample solution and reading the result through the instrument. However, QDs and CNTs have also been employed to detect MC. QDs, a part providing unique fluorescence properties, have also been exploited for the development of electrochemical devices due to their properties as semiconductors. Consequently, Yu et al. [132] have prepared an electrochemical immunoassay for the detection of MC-LR. The assay was performed using a quantum dots/antibody (QDs/Ab) probe to convert the biorecognition event in an electrochemical signal by measuring the cadmium ions released from QDs based on Square Wave Stripping Voltammetry (SWSV). The QD/Ab

2 New techniques in environment monitoring

T

B

Inavalid

5

Positive (+)

4

Positive ( + –)

3

Positive ( + –)

2

Negative (–)

1

|

63

C

T

C

IP VC

pl at e

III Gold conjugate pad Flow

II Sample pad A

IV NC membrance

V Absorbent pad

Anti-OA McAb

OA conjugate

Colloidal gold-labeled anti-OA McAb

Goat anti-mouse lgG

CT

Fig. 2.2: A – Schematic diagram of the test principle of LFIC for OA. Details are included in the text. B – Estimate of the LFIC test. C, control line; T, test line. Strip 1 indicates negative results when the color density of the test line is the same as that of the control line. Strips 2 to 4 indicate positive results when the color density of the test line is weaker than that of the control line (strips 2 and 3) or when the concentration of the toxin is high enough to neutralize all colloidal gold-conjugated McAb, resulting in nocolor (strip 4). Strip 5 indicates an invalid test result when no lines appear in the control line. Reprinted from, Analytical Biochemistry 422, Lu SY, Lin C, Li YS, Zhou Y, Meng XM, Yu SY, Li ZH, Li L, Ren HL, Liu ZS, A screening lateral flow immunochromatographic assay for on-site detection of okadaic acid in shellfish products, 59–65, Copyright (2012), with permission from Elsevier.

1 2 3 4 5 6 7 8 9

Fig. 2.3: LFIC assay for a series of OA dilutions using gold particle probes. The assay was carried out by applying 50 ll of the standard toxin solution at different concentrations with an equal volume of sample buffer to the wells. C, control line; T, test line. Strips 1 to 9 have the following OA concentrations (from left to right): 0, 1.25, 2.5, 5, 10, 20, 30, 50, and 70 ng ml−1 . Reprinted from, Analytical Biochemistry 422, Lu SY, Lin C, Li YS, Zhou Y, Meng XM, Yu SY, Li ZH, Li L, Ren HL, Liu ZS, A screening lateral flow immunochromatographic assay for on-site detection of okadaic acid in shellfish products, 59–65, Copyright (2012), with permission from Elsevier.

64 | Begoña Espiña et al.

probe showed remarkably high sensitivity in a range of 0.227–50 μg/l with a low LOD of 0.099 μg/l. Furthermore, the direct transduction of the electronic signal could provide flexibility for developing a biosensor format with the advantages of mass production of a miniaturized and portable device. Wang et al. have used SWCNT to develop a biosensor based on a large surface area and the electrical conductivity of nanomaterials [133]. The approach consisted of impregnating SWCNT with antibodies on regular paper strips. The dense nanotube percolation network favors the antibody-antigen complex. Therefore, the presence of the analyte alters the current flowing through the electrode, and the conductivity based on the SWCNT-SWCNT charge transfer was strongly dependent on analyte concentrations. This method showed a LOD of 0.6 nmol/l, i.e. 0.6 ng/ml; this satisfies the WHO standard for MC-LR content in drinking water (1 ng/ml). The time of detection was about 28 times less than the conventional ELISA; this is a significant reduction in analysis time. A graphene-based immunosensor was fabricated for the detection of MC-LR using a multienzyme-carbon nanosphere-antibody system for signal amplification. This approach provided a linear range from 0.05 to 15 μg/l of MC-LR with a detection limit of 0.016 μg/l [134]. Another example of electrochemical immunosensors for MC-LR detection is presented by Zhang et al.: single-walled carbon nanohorns (SWNHs) were used for analyte immobilization. Compared with single-walled carbon nanotubes, SWNHs as immobilization matrixes showed a better sensitizing effect. Using an enzyme-labeled MC-LR antibody for the competitive immunoassay, under optimal conditions, the immunosensor exhibited a wide linear response to MC-LR ranging from 0.05 to 20 ng/ml with a LOD of 0.03 ng/ml. The proposed strategy provided a biocompatible immobilization and sensitized recognition platform for analytes such as small antigens and possessed promising application in food and environmental monitoring [135]. Further improvement on sensitivity (LOD of 3.7×10−17 M ≈ 37 fg/l), repeatability and stability when compared to previous MC-LR sensors was achieved using hybrid graphene-gold nanocomposite film with electrodeposition [136].

2.6 Molecular biology-based techniques for HABs detection 2.6.1 Overview Currently, the monitoring of microalgae is mainly based on morphology determined by light microscopy. However, such approaches – along with being time consuming – do not provide sufficient information, especially concerning species and toxin attribution, and it requires a high degree of expertise on behalf of the operator [137]. Speciesspecific identification of cells through this approach appears to be specially cumbersome for some cryptic species such as Pseudo-nitzschia spp., or those with difficult

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morphotype identification such as Alexandrium spp., Chattonela spp. or Ostreopsis spp. [1]. A very illustrative case of such difficulties was the identification of Alexandrium minutum as the PSP-producing species in Cork Harbour, Ireland with the advent of molecular biology-based techniques [138], while for over a decade the production of PSP was attributed to Alexandrium tamarense [41]. Since data derived from monitoring programs is used to produce predictive models in order to forecast toxic blooms [137], the accuracy of identification methods for toxic algae is of high importance. Another drawback for microscopy techniques is the long time needed for analysis. This fact induces the requirement of preservation procedures in order to maintain cell integrity through the entire analytical procedure, and at the same time compromises the early detection of HABs [1] and therefore a rapid response in order to enable the prevention and/or mitigation of these episodes on the ecosystem. Molecular biology techniques have been pointed out as being a good alternative for the monitoring of toxic algae. Since such techniques are usually faster and more reliable, they provide an increased accuracy when compared with light microscopy, and they allow the identification to the species level without the need of personnel trained in taxonomy [137]. These methods also contribute to understanding the presence of toxic microalgae and their distribution and dispersion mechanisms, facilitating the prevention or mitigation of their effects both on human health and marine and freshwater ecosystems and therefore on related economic activities [1]. Molecular biology-based techniques offer a near real-time analysis of the ecosystem and, at the same time, they offer broader ecological interpretation of how key species, such as toxic algae, can extend their distribution with climate change to become invasive species [139]. In the last several years, numerous research groups have made use of identification methods based on nucleic acid sequences available on public databases and thus developed analytical approaches based on molecular methodologies in order to identify and quantify certain species and the toxins they produce. For example, cyanobacterial blooms may contain both toxic strains, able to produce MC as well as nontoxic ones, and the only way to distinguish between the two types of strains is to apply molecular biology-based methods in order to estimate the presence of MC synthetase (mcy) genes, which are responsible for the production of MC [140]. Among these approaches we can find two main groups: those based on hybridization-based techniques such as fluorescence in situ hybridization (FISH), SHA or microarrays, and amplification-based techniques such as end-point and quantitative polymerase chain reaction (qPCR) or loop-mediated isothermal amplification (LAMP).

2.6.2 DNA/RNA targets One of the key factors for the development of such methods is the choice of an appropriate target. Commonly used DNA targets for the identification of specific organisms

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in environmental monitoring include single copy or multiple copy targets such as ribosomal, chloroplast, mitochondrial or other highly repetitive sequences for eukaryotic species. Availability of DNA/RNA sequences on public databases is a limiting factor for the development of molecular biology-based methods; fortunately, in recent years advances in next-generation sequencing (NGS) have increased the number of available sequences on public databases, reduced the cost and simplified the process for the sequencing of genes of interest by research groups [141]. Two main groups of target DNA sequences are being used for the development of methods in this area: (1) sequences that allow the specific identification of one particular or group of toxic algae; (2) genes involved in the production of their toxins. The choice of one approach over another approach should be directed by analytical needs. Among the first group, ribosomal DNA (rDNA) regions have been frequently used as targets for molecular biology-based methods of analysis for different organisms. Such regions are repeated in tandem and high copy numbers which contributes to improving the sensitivity of such methods; at the same time, they are phylogenetically informative [1]. Moreover, a high number of sequences are available in public databases, which facilitates probe design and comparison with other related organisms. Among frequently used regions of the rDNA cluster, we can find the small subunit (SSU), the large subunit (LSU), the 5.8S region, two internal transcribed spacers (ITS1, ITS2) and the non-transcribed spacer (NTS) as seen in Tab. 2.1. Another potential target is the genes that encode the carbon fixation enzyme ribulose-1,5-biphosphate carboxylase/oxygenase (RuBisCO), which consists of eight large and eight small sub-units encoded by the rbcL and rbcS genes. These genes are chloroplast-encoded and cotranscribed [142]. Both rbcL and rbcS have been suggested as potential targets for species identification due to their diversity; this allows for the resolution of closely-related species [143]. Concerning the second group of DNA targets, their use is highly influenced by the identification of the genes involved in the production of toxins and the availability of their specific sequences. These targets are particularly attractive for environmental monitoring, since they allow one to establish the potential for production of such toxins, and to investigate which environmental factors influence their production. Among harmful algae produced toxins, STX-producing genes and pathways have been identified for both dinoflagellates [144, 145] and cyanobacterial species [145– 149]; other described cyanotoxin-producing gene clusters including MCs (mcyS), nodularin (ndaS) and cylindrospermosin (cyr) [150–152], as well the genes involved in the anatoxin-a and homoanatoxin-a, have been recently identified [153]. Together with DNA targets, gene transcripts (mRNA) have also been used for the detection of several toxic microalgae by amplification-based methods [143, 154–157]. Such methods enable a more accurate estimation of potential toxin-producing cell concentration, since the fast degradation of mRNA only allows for the detection of viable cells, and a high sensitivity due to both the amplification and the presence of high

2 New techniques in environment monitoring |

Tab. 2.1: Molecular methods. Target Organism

Target genomic region

Method

Reference

Cyanobacteria

16S rRNA

qPCR

[220, 242]

Cyanobacteria

16S rRNA

End-point PCR and qPCR

[243]

Cyanobacteria

mcyA gene

PCR-DGGE and qPCR

[244]

Cyanobacteria

mcyE/ndaF genes

PCR-DGGE and qPCR

[245]

Microcystis spp.

mcyB gene

End-point PCR

[212]

Microcystis spp.

mcyB gene/PC gene region

qPCR

[246]

Microcystis spp.

16S rRNA gene/mcyA and mcyB gene

qPCR

[220]

Microcystis spp.

16S rRNA/mcyB and mcyD genes

End-point PCR and qPCR

[197, 211, 242, 243]

Microcystis spp.

16S rRNA/mcyD gene

qPCR

[215]

Microcystis spp.

16S rRNA/mcyA

qPCR

[210]

Microcystis spp.

rRNA internal transcribed spacer region

PCR-DGGE

[217]

Microcystis aeruginosa

PC gene/mcyA

qPCR

[218, 247]

Microcystis spp. and Anabaena spp.

mcyE

qPCR

[214]

Microcystis spp.

PC gene/mcyB

qPCR

[248, 249]

Microcystis

mcyA

qPCR

[250]

Planktothrix

PC gene/mcyA

qPCR

[246, 251]

Species and genera belonging to Dinophyceae, Bacillariophyceae and Raphydophyceae.

ITS-5.8S rDNA

End-point PCR

[252]

A. minutum

ITS-5.8S

qPCR

[181, 201]

A. catenella

LSU (D1/D2)

qPCR

[195, 200, 253]

A. catenella

ITS-5.8S

qPCR

[187]

A. catenella

sxtA4

qPCR

[203]

Cyanobacteria

Dinoflagellates

67

68 | Begoña Espiña et al. Tab. 2.1 (continued) Target Organism

Target genomic region

Method

Reference

A. catenella and A. minutum

ITS

LAMP

[221]

Alexandrium spp.

5.8S

LAMP

[222]

A. fundyense

LSU

qPCR

[199, 254]

A. tamarense

LSU

qPCR

[200, 253, 255]

A. taylori

ITS-5.8S

qPCR

[187]

Cochlodinium polykrikoides

LSU

qPCR

[206]

Cochlodinium polykrikoides and C. brodyi

ITS

qPCR

[256, 257]

Dinophysis acuta and D. acuminata

LSU (D1/D2)

qPCR

[190]

Gambierdiscus belizeanus, G. caribaeus, G. carpenter, G. carolinianus, G. rutzleri

LSU, SSU, 5.8S

qPCR

[258]

Heterocapsa circulasquama

LSU (D1/D2)

qPCR

[206]

Lingulodinium polyedrum

SSU

qPCR

[259]

Karenia brevis

rbcL transcript

NASBA

[154]

K. brevis

rbcL Transcript

qRT-PCR

[156]

Karenia mikimotoi

rbcL Transcript

NASBA

[157]

K. mikimotoi

LSU

qPCR

[206]

K. mikimotoi

ITS-5.8S

qPCR

[260]

Ostreopsis cf. ovata

LSU (D1/D2)

qPCR

[186]

Pfiesteria piscida

mt Cob, SSU

qPCR

[261]

P. piscicida

cyt c oxidase subunit 1, cyt b (COB), and Tags 343 and 277

qRT-PCR

[155]

P. shumwayae

mt Cob, SSU

qPCR

[262]

Chattonella spp., Heterosigma akashiwo

LSU (D1/D2)

qPCR

[263]

C. verrucolosa, C. antiqua, C. ovata, C. subsalsa, Fibrocaspa japonica

SSU/ITS2

qPCR

[204]

C. cf. verrucolosa, C. subsalsa, Heterosigma akashiwo

SSU

qPCR

[205]

Raphydophyceae

2 New techniques in environment monitoring |

Tab. 2.1 (continued) Target Organism

Target genomic region

Method

Reference

ITS2

qPCR

[207, 208]

SSU

qPCR

[209]

Pseudo-nitzschia spp.

SSU

qPCR

[194]

Pseudo-nitzschia spp

ITS-5.8S

qPCR

[193]

Pseudo-nitzschia multiseries

rbcL transcript

NASBA, qRT-PCR

[143]

A. fundyense, P. australis, and A. ostenfeldii

ribosomal RNA (rRNA)

Microarray

[264]

Eukaryotes, Chlorophyta, Bolidophyceae, Prymnesiophyta and Cryptomonads

SSU rRNA

Microarray

[265]

Alexandrium tamarense, A. catenella, A. andersoni, A. pseudogoniaulax, A. minutum, A. taylori, L. polyedrum, Protoceratium reticulatum, Dinophyceae

ITS1-5.8S-ITS2 rDNA regions

Microarray

[266]

Alexandrium spp.

18S rDNA

Microarray

[169]

Ale. catenella, Ale. minutum, Ale. tamarense, Coc. polykrikoides, G. aureolum, G. catenatum, G. impudicum, Aka. sanguinea, H. akashiwo, Cha. marina

LSU rDNA D2 regions

Low-density oligonucleotide array

[170]

Pseudo-nitzschia species

Nuclear LSU and SSU rRNA

Microarray

[159]

Eukaryotes, Cryptophytes, Cryptosporidium, C. parvum, Navicula accomoda, Nitzschia communis

18S rRNA genes

Microarray

[171]

Pseudo-nitzschia spp.

ITS1 region

Microarray

[267]

Haptophyceae Prymnesium parvum Pelagophyceae Aerococcus anaphagefferens Diatoms

69

70 | Begoña Espiña et al. Tab. 2.1 (continued) Target Organism

Target genomic region

Method

Reference

A. minutum, A. ostenfeldii, D. acuminata, D. acuta, Heterosigma akashiwo, Karenia brevis, Karenia mikimotoi, Karlodinium veneficum, Pseudochattonella verruculosa, Pseudochattonella farcimen, Pseudo-nitzschia australis, Pseudo-nitzschia calliantha, Pseudo-nitzschia multiseries, Pseudo-nitzschia multistriata, Prymnesium parvum, Prymnesium (0Chrysochromulina) polylepis

18S and 18S RNA

Microarray

[139, 172–175]

qPCR: quantitative PCR, PCR-DGGE: PCR-Denaturing Gradient Gel Electrophoresis, LAMP: Loop Mediated Isothermal Amplification, qRT-PCR: quantitative reverse transcription PCR, NASBA: Nucleic acid sequence-based amplification. PC: Phycocyanin, ndaF : nodularin synthetase gene subunit F, mcy: microcystin synthetase, ITS: Internal Transcribed Spacer, LSU: large subunit RNA, SSU: Small Subunit RNA, sxt: saxitoxin, cyt: cytochrome, rbcL: RuBisCO large subunit.

copy numbers of mRNA per cell [143]. Amplification methods using mRNA include quantitative reverse transcription PCR (qRT-PCR) and nucleic acid sequence-based amplification (NASBA) as described below. rRNA is frequently used as a target for hybridization based techniques [158, 159]; these molecules are found in high target numbers in cells, and the design of probes at different taxonomic levels is possible due to the use or more-or-less conserved regions [158, 160].

2.6.3 Hybridization-based techniques Hybridization-based techniques rely on the use of molecular probes to identify species of interest by binding to the target’s sequence and being subsequently detected by a probe-attached label, e.g. digoxigenin or a fluorochrome such as fluorescein [137].

2.6.3.1 FISH and SHA In FISH, a fluorescently labeled probe designed to recognize a specific sequence of a particular organism is hybridized inside intact cells, and cells containing a complex

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of the probe and the specific sequence are detected using epifluorescence microscopy. This technique does not require cell homogenization or lysis and allows the analysis of whole cells by their morphological characteristics [161]. In contrast, SHA relies on the lysis of cells to release nucleic acids; visualization of the microalgae under a light microscope is not required [162]. This approach uses two oligonucleotide probes that typically target ribosomal RNA (rRNA). One probe captures the target molecule and the other probe, also binding in close proximity to the capture probe, carries a signal moiety that can be detected by various means. For available methods developed for toxic microalgae, oligonucleotide RNA probes have a length of 18–25 base pairs and target complementary sequences of the small and large sub-unit ribosomal RNAs. Both FISH and SHA assays have been developed and applied to the identification of toxic microalgae in recent years [158, 162–166]; however, according to Rhodes et al., qPCR methods appear to be a more rapid and sensitive approach than FISH or SHA assays for some of the toxic microalgae such as the species in the Karenia genus [162, 167]. Both hybridization-based techniques have a few drawbacks which have led to an increase in the use of amplification-based techniques. As an illustrative example, SHA has been widely used for research purposes, but has not proven to be simple and fast enough to replace current methods [167]. Moreover, the sensitivity of SHA appears to be insufficient to meet the monitoring requirements for some species such as the STX-producing Alexandrium spp. (i.e. a trigger level of 100 cells/l).

2.6.3.2 Microarrays Microarray is another hybridization-based technique in which specific RNA/DNA sequence probes are spotted on a surface, usually a glass microscope slide. Each of these spots is complementary to an extracted target (RNA or DNA). A fluorescent label allows one to measure the amount of target in the sample using a microarray scanner [168]. Microarrays offer the possibility to detect multiple species simultaneously using species-specific probes, can give a broad range of information concerning target HAB species in a single experiment and offer the possibility of obtaining a high sample throughput [139]. Most of the published work so far has relied on a prior amplification by PCR of the target sequence from the species of interest with universal primers, and on the use of specific immobilized probes between 18 and 25 nucleotides [169–171]. However, there are also some works in recent years that have targeted the direct use of extracted DNA/RNA sequences, such as the case of some methods developed under the EU FP7 project: microarrays for the detection of toxic algae (MIDTAL), in which rapid species identification using rRNA genes was developed for use in a species microarray or phylochip [137]. Some of the systems developed under the framework of the MIDTAL project were based on the use of extracted RNA, without an amplification

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step; this approach has several advantages, including the avoidance of the PCR step, and that the presence of the rRNA target can indicate bioactive plankton cytoplasm since it degrades very easily in the environment [159]. The use of microarray technology as a routine monitoring tool for early warning systems for toxic algae needs some improvements. The use of microarrays for toxic algal species has been evaluated and compared with cell counts to validate their potential to monitor the Spanish Atlantic coast (NW Iberian Peninsula, Galicia); it offers a good correlation and provides better species resolution in Alexandrium and Pseudonitzschia [139]. Likewise, the microarray was tested on Arcachon bay (France) [172], in Irish coastal waters [173] and in the Orkney Islands, UK [174] with similar results. The MIDTAL microarray has also been compared with qPCR and microscopy for the monitoring of seasonal dynamics of harmful algae in outer Oslofjorden, southern Norway [175].

2.6.4 Amplification-based techniques 2.6.4.1 End-point PCR and qPCR PCR is the most commonly used method for the amplification of specific DNA sequences in clinical diagnosis, food analysis and environmental monitoring. With PCR, a specific sequence within a DNA template can be amplified up to a million times [176]. The specificity of the methods is achieved by the use of primers that recognize DNA fragments from the organism/organisms of interest. It has been successfully applied to the detection of various toxic microalgae in seawater and freshwaters samples. It has interesting advantages when compared to other molecular techniques, including the amplifying of the number of copies of the initial DNA target, which enables higher sensitivity, and the possibility of quantification by the use of qPCR. End-point or classical PCR is the simplest and probably least expensive approach for DNA amplification; its main drawbacks are that this PCR method requires an additional step for the detection of the amplification product, which traditionally has been performed by gel electrophoresis, and that this method is a merely qualitative technique. The quantitative RT-PCR or qPCR allows DNA amplification to be monitored while it is being produced because the signals (generally fluorescent) can be observed as they are being generated during the amplification of the target DNA [177]. As the amount of product accumulates, the signal increases exponentially, then the signal levels off and saturates because critical components of the reaction run out [178]. Initial template levels can be calculated by analyzing the shape of the curve or by determining when the signal rises above a certain threshold value. The observed increase in fluorescence is proportional to the starting amount of target molecules. qPCR methods typically make use of two detection chemistries. The first is a DNA binding dye, most commonly SYBR-Green [179], which is a cheaper alternative. However, there

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are some drawbacks, such as its lack of specificity to bind double-stranded DNA. The second is a hydrolysis probe approach, which usually uses TaqMan® Probes that increase the specificity of the assays. With this methodology, a positive identification requires the effective binding of a specific probe in addition to the binding of the PCR primers. Simultaneously, qPCR allows the expression of the results as numerical values (Cq values), which provides more information about the kinetics of the reaction. Quantification cycle (Cq) values are calculated by determining the point at which the fluorescence that is produced in each sample reaches a set threshold limit. This situation occurs during the exponential phase of the reaction and, therefore, before some of the reagents are consumed because of amplification, which would negatively affect the reliability of the quantification results – at this point, the reactions start to slow down, and the PCR product is no longer being doubled at each cycle. Cq values are inversely related to the starting copy number of the target sequence [180]. Both PCR and qPCR can be applied to seawater, sediments and seafood animals that have incorporated the toxic species into their body by filtration [181, 182]. Compared to end-point PCR, qPCR presents several advantages, including the possibility of using shorter fragments. Because there is no need for gel visualization [183], this technique is applicable to difficult samples such as highly processed samples, which may have highly fragmented DNA, reporting better results than end-point PCR [184]. QPCR is an extremely sensitive molecular method, and in recent years it has been developed and extensively applied to the identification and estimation of toxic phytoplankton species in natural marine samples [143, 185, 186], as recently reviewed by Penna and Galluzi [1], and in freshwater. Advances in the development of qPCR for specific applications in this field have had a very positive influence on the advancement in the study of microalgal population dynamics in marine ecosystems, since it allows both the identification and enumeration of target phytoplankton species in field samples [181]. As mentioned above, qPCR has a quantitative potential; however, the accuracy of the estimation of DNA copy numbers and its relationship with cell abundance might be affected by several factors that must be carefully evaluated in order to obtain reliable results [187, 188]. Most published methods in this area are based on an absolute quantification approach. Such methodology relies on the use of a standard curve that is constructed by amplifying known amounts of target DNA in a parallel set of reactions. Such standard curves are frequently produced using a plasmid containing the cloned target sequence [181, 189, 190], from serial dilutions of DNA extracted from clonal cultures [191–193] or from lysates of target cells that have been previously serially diluted and spiked into natural seawater or medium [194, 195]. An absolute quantification approach requires on the one hand that both the known DNA (from the standard curve) and the investigated DNA (from the sample) are amplified with the same efficiency [196] in order to obtain reliable quantitative results. On the other hand, for the estimation of cell content in a particular sample, it

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is important that the amount of target gene per cell is known [1], different gene copy numbers per cell at different stages of the microalgae growth might affect quantitative results [197]. Despite these possible fluctuations in quantitative results, qPCR still offers great potential for the estimation of the risks associated with HABs and moreover how environmental factors can affect the growth of toxic algae and the production of associated toxins. Methods based on qPCR have been developed for several microalgae, such as dinoflagellates, specially Alexandrium spp. [181, 187, 195, 198–203], from the Raphidophyceae class such as Chattonella spp. [204–206], from Haptophytaceae [207, 208], Pelagophyceae [209], and diatoms, in particular Pseudo-nitzschia spp. [143, 193, 194]. QPCR is also becoming widely used in monitoring toxic cyanobacterial blooms; several groups have used this technique in order to estimate the number of toxic genotypes and their proportions in the total pool of cyanobacteria [176, 192, 210–214]. PCR-based methods are contributing to a better understanding of how environmental factors can affect quantities of toxic cyanobacteria, and therefore the concentration of MCs [215–220]. Quantitative reverse transcription PCR (qRT-PCR) is a variant of qPCR in which complementary DNA (cDNA) is produced from transcripts from RNA and enables the study of gene expression. This technique has also been used to evaluate mRNA detection capabilities for the identification of viable cysts of Pfiesteria piscicida in natural sediment samples [155], for the detection and quantitation of the Florida red tide organism, Karenia brevis [156], and for the detection of the toxic species Pseudo-nitzschia multiseries [143].

2.6.4.2 New amplification strategies New amplification strategies are being developed in order to facilitate simpler and faster detection and identification of toxic algae. Recently, a method based on LAMP was designed and evaluated for rapid detection of the toxic microalgae Alexandrium catenella and A. minutum [221] and for the detection of Alexandrium species with the use of a set of four primers targeting six conserved sequences of the 5.8S region [222]. LAMP techniques allow for the amplification of a few copies of a target DNA sequence to 109 copies under isothermal conditions (in the range of 60–65 °C) in less than an hour [223]. The LAMP technique requires a set of four specific primers recognizing six distinct sequences on the target DNA, Bst DNA polymerase and a heat block or water bath, which make it simpler to perform than PCR, which requires specific equipment. In this study, the detection of positive samples was possible by visual inspection thanks to either a white precipitate due to the presence of magnesium pyrophosphate or a color change due to the addition of intercalating SYBR-Green I, which also contributes to the ease of use of such techniques for fast and sensitive detection of these organisms [221].

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NASBA is also an isothermal method (41 °C) for RNA amplification that is catalyzed by an enzyme mix consisting of T7 RNA polymerase, avian myoblastosis virus reverse transcriptase, and RNaseH and two target-specific oligonucleotide primers [157]. Quantification of RNA by real-time NASBA is accomplished by the use of a molecular beacon labeled with a fluorophore, a single-stranded oligonucleotide that forms a stem-loop structure [224]. During the amplification reactions, the fluorescent signal is measured and quantification of RNA occurs by comparing time to positive (TTP) values (analogous to threshold cycle in qPCR) of fluorescence amplification plots to TTP values of a known standard curve [157]. Methods based on this technique have been developed for the detection of K. brevis [154], detection and quantification of Karenia mikimotoi [157] and detection of Pseudo-nitzschia multiseries [143].

2.6.5 Aptamers for toxin detection Another interesting trend in the environmental monitoring of marine toxins and their microalgae is the development of more specific and stable ligands, which might greatly contribute to the development of more reliable detection methods. In this sense, aptamers are gaining the attention of some research groups. Aptamers are single stranded (ssDNA) or RNA oligonucleotides that can bind diverse targets, such as small ions, proteins or even cells. Aptamers possess a high affinity for their targets due to their capability to fold upon binding with their target molecule [225]. Aptamers are selected in vitro from large populations of random sequences through a combinatorial strategy called SELEX (systematic evolution of ligands by exponential enrichment) [226]. Aptamers are a relevant alternative to antibodies in bio-analytical applications [227] due to their advantages when compared with antibodies, such as: (1) the in vitro selection procedure and chemical synthesis of aptamers, which eliminates the need for the in vivo immunization of animals to generate antibodies; (2) the possibility of selecting a new range of targets, including toxic or non-immunogenic materials; (3) no batch-to-batch differences in activity; (4) cost-effective synthesis; and (5) aptamers can be easily modified with a variety of functional groups or molecules that facilitate the immobilization of aptamers on a solid support or the introduction of a label [228], which is of high importance when we want to develop faster, but reliable, analytical methods such as biosensors. In recent years, several aptamers have been selected for the detection of marine and freshwater toxins; DNA aptamers with a high affinity and specificity to MC-LR, -YR, and -LA have been selected and used on electrochemical aptasensors which provided congener-specific MC detection and a detection limit of 10 pM [229]. Recently, a previously selected aptamer for MC-LR [230] was used on a label-free aptamer-based electrochemical impedance biosensor which showed a good linear relationship in the range of 1.0 × 10−7 to 5.0 × 10−11 mol/l when plotting the measure of the decrease of

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the impedance rate with the logarithm of the MC-LR concentration and a calculated LOD of 1.8 × 10−11 mol/l [231]. A DNA aptamer for okadaic acid was also selected and used for the fabrication of a label-free electrochemical biosensor [232] showing a linear range for the concentrations of OA between 100 pg/ml and 60 ng/ml with a LOD of 70 pg/ml. A DNA-aptamer targeting STX was also selected; concentration-dependent and selective binding of the aptamer to STX was determined using a surface plasmon resonance sensor [233]. Hu et al., selected a DNA aptamer against fragments of STX antibodies from a random ssDNA library by SELEX; the selected aptamer showed a dissociation constant (Kd) in the nanomolar range [234]. An interesting application of aptamers on environmental control was proposed by Hu et al. in which an aptamer (RNA) specific for MC-LR was covalently immobilized on graphene oxide to form a polydispersed and stable nanosheet and used as a novel sorbent for MC-LR in drinking water [235].

2.7 Future perspectives Further development of all of the above-reviewed technologies will help us to better comprehend how HABs’ occurence, frequency and severity are related to the changes induced by so-called climate change. New trends in environmental monitoring try to make the collection of samples and analysis simpler, multiplex, on-the-spot and automated when possible. The increasing interest, development and investment in nanotechnology and biotechnology make this field one of the most promising approaches for improving HABs monitoring systems. The development of new customized nanomaterials as absorbents in SPME for toxin extraction such as Molecular Imprinted Polymers and Carbon Nanotubes is of high interest due to their efficiency and selectivity. Porous nanomaterials such as Organic Frameworks (OF) are currently being evaluated in the International Iberian Nanotechnology Laboratory for their use in applications where waterborne toxins should be adsorbed: in detoxification, separation and monitoring. Those materials have much more restricted pore sizes than the commonly used resins for SPATT or chromatographic separation. Additionally, OFs can be decorated (functionalized) with diverse groups; this changes their affinity to different molecules and makes them excellent candidates to selectively separate and concentrate toxins depending on their nature and size. Space-resolved SPME uses miniaturized segmented fibers, 4–5 mm apart, to allow for simultaneous sampling of two different tissues or stratified matrixes. SR-SPME is simpler and more cost-effective [31] and could be used in future applications to sample seafood for safety monitoring. This would significantly simplify the sample preparation process prior to analysis.

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The trend of biosensing technologies for HABs and their toxins has experienced a paradigm shift away from conventional to molecular and nanotech-based assay techniques. Although conventional methods are still recognized as standard techniques for the identification of toxins, they are laborious and time-consuming. The new trend is to invest in advanced detection tools allowing for the fabrication of rapid and userfriendly advanced biosensor devices imperative for rapid, continuous, in situ toxin detection and identification. However, a main concern/limitation common to all biosensors is the bioreceptor. So far neither aptamers nor MIPs are competitive with antibody-based bioreceptors. However, problems can even occur with antibody-based biosensors, related to cross-reactivities and specificities which highly influence the sensitivity of the sensor [236]. Recent advances report label-free sensors combined with cell-based detection systems for toxins; they are able to detect low levels of multiple contaminants simultaneously in an early stage of the bloom [236, 237]. Future improvement in the development of nanoparticles with enhancement of their physical and chemical properties will contribute to producing more effective biosensors. Thus environmental monitoring would profit from the new discoveries. Multidetection analysis, higher sensitivities and miniaturization are some of the goals in biosensing that nanotechnology researchers should focus on. In this sense, subnanometric materials could play an important role. As an example, metal atomic clusters consist of groups of atoms with well-defined compositions that represent (after atoms) the most elemental building blocks in nature. Atomic clusters, characterized by their extremely small size, comparable to the Fermi wavelength of an electron, makes them a bridge between atoms and nanoparticles or bulk metals, with properties very different from both of them [238]. Small metal clusters have been found to display photoluminescence properties, not found in bulk or nanoparticles. The appearance of optical transitions with energies ranging from the UV-visible to the infrared region makes noble metal clusters ideal potential candidates for applications such as fluorescent labeling in biological media. In the NANOMAG group of the University of Santiago de Compostela (NANOMAG), fluorescent systems were developed that were based in stable metallic clusters (⪅ 2 nm) with promising results for optical sensing [239, 240]. In regards to new molecular biology approaches, the tendency nowadays is to substitute the LM for molecular biology-based techniques in order to have a faster and more specific identification of toxic microalgae. The shift has been especially towards rapid throughput qPCR assays [167] and microarrays [137]. In addition to this trend towards faster and more specific methodology, multi-species detection systems are also highly required in order to be able to monitor several toxic microalgae at the same time. In this sense, approaches being developed in the microarray field are very positive in order to develop multiplexed analytical procedures. The use of high resolution sampling combined with molecular biology-based techniques can greatly contribute to better environmental monitoring [241]. Integration of sampling and analysis through molecular biology-based techniques such as SHA

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[166, 241] or qPCR [8] has been applied recently, and such an approach could lead to an important improvement in actual tools for environmental monitoring. Other interesting developments with applications in the field are the improvements made in next-generation DNA sequencing technology. This will contribute to finding more apppropiate targets for the specific identification of toxic microalgae; at the same time, it is possible that handheld sequencers will be the future of phytoplankton monitoring programmes [167]. All these tools have the potential to support predictive modelling, which is being developed in order to give warning of an increased risk of an occurring event [167].

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Abbreviations ABDMAP AOBS A-OVA ASP AuNPs AuNRs AZP BSA BTX cDNA CLIA CL-MADAG CPRS CFP CTXs Cq DA DAD DNA DPV DSP DTX-1 ELISA

Algal Bloom Detection, Monitoring And Prediction automated online optical biosensing system antigen-ovoalbumin amnesic shellfish poisoning gold nanoparticles gold nanorods azaspiracid bovine serum albumin brevetoxin complementary DNA ChemiLuminscence ImmunoAssay chemiluminescence multichannel immunosensor Continuous Plankton Recorder Surveys Ciguatera Fish Poisoning ciguatoxins Quantification cycle domoic acid Diode Array Detector deoxyribonucleic acid differential pulse voltammetry diarrheic shellfish poisoning dinophysistoxin-1 enzyme-linked immunosorbent assay

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EO EPA ESP FDA FISH GACS GGNRS GO GTX 2/3 HPLC HPLC-MS/MS HP-LFIA HRMS HRP IC20 IC50 IFCB IgG IPCC IT IUPAC LAMP LC-MS/MS LFIA LFIC LOD LPB LSPR LSU mAb MBA MC MERHAB MR mRNA NASBA NERSC NGS NOAA NPs NSP NTS OA OLCI PEF PEI PEI-MBs PEMC POCISs

Earth Observation US Environmental Protection Agency Environmental Sample Processor US Food and Drug Administration fluorescence in situ hybridization Global Alliance of Continuous Plankton Recorder Surveys guanine-assembled graphene nanoribbons graphene oxide gonyautoxin 2/3 High Performance Liquid Chromatography high performance liquid chromatography coupled to tandem mass spectrometry high performance lateral flow immunoassay High Resolution Mass Spectrometry HorseRadish Peroxidase Inhibitory Concentration 20 inhibitory concentration 50 Imaging FlowCytobot immunoglobulin G Intergovernmental Panel on Climate Change ion trap mass analyzers International Union of Pure and Applied Chemistry loop-mediated isothermal amplification Liquid Chromatography coupled to tandem Mass Spectrometry Lateral Flow Immunoassay lateral flow immunochromatographic limit of detection longitudinal plasmon band localized surface plasmon resonance large subunit monoclonal antibody Mouse BioAssay microcystin Monitoring and Event Response for Harmful Algal Blooms magnetoresistive messenger RiboNucleic Acid nucleic acid sequence-based amplification Nansen Environmental and Remote Sensing Center next generation sequencing National Oceanic and Atmospheric Administration nanoparticles Neurotoxic Shellfish Poisoning nontranscribed spacer okadaic acid Ocean Land Colour Instrument pefloxacin polyethyleneimine PEI-modified magnetic beads piezoelectric-excited millimeter-sized cantilevers Polar organic chemical integrative samplers

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PSP PZT QCM QDs qPCR QqQ qRT-PCR rDNA RuBisCO SAHFOS SELEX SHA SPATT SPME SPR SRM ssDNA SST SSU STX SV SWCNT SWNHs ToF TTP TTX UPLC WHO μ-TAS

paralytic shellfish poisoning lead zirconate titanate quartz crystal microbalance quantum dots quantitative Polymerase Chain Reaction triple-quadrupole mass analyzers quantitative reverse transcription PCR ribosomal DNA ribulose-1,5-biphosphate carboxylase/oxygenase Sir Alister Hardy Foundation for Ocean Science systematic evolution of ligands by exponential enrichment Sandwich Hybridization Assay Solid-Phase Adsorption Toxin Tracking Solid Phase Micro-Extraction surface plasmon resonance selected reaction monitoring mode single stranded Sea Surface Temperature small subunit saxitoxin Spin valves single walled carbon nano tubes single-walled carbon nanohorns time-of-flight mass analyzers time to positive tetrodotoxin Ultra Performance Liquid Chromatography world health organization micro-total analysis system

Mikko Nikinmaa and Katja Anttila

3 Responses of marine animals to ocean acidification 3.1 Introduction The acidification of oceans due to increased atmospheric carbon dioxide levels is clear and progressing. Oceanic pH has decreased from the pre-industrial value of approximately 8.2 to the present value of approximately 8.1 [1]. The predictions propose an overall pH decrease to accelerate with a maximal 0.5-unit decrease by 2100, although most predictions estimate more modest changes of 0.3–0.4 units [2]. Moreover, ocean acidification is usually considered together with climate change, as anthropogenic causes are considered to be behind both. It is often thought that both would necessarily be interconnected. Notably, ocean acidification can occur even if it is thought that human-induced effects do not cause temperature changes. However, as described below, there are intimate interactions between the temperature change and ocean acidification. In this chapter we first evaluate what determines the acid-base balance of the marine environment, and what causes its pH perturbations. In the overview, it should be borne in mind that there are many reviews that discuss the mechanisms behind ocean acidification much more thoroughly than our treatment. Thereafter, we discuss the effects that acid-base disturbances in marine environments can have on animals. The first of these is organismic pH regulation; here all the effects associated with calcification are treated. We also focus on possible developmental effects and consider respiratory and behavioral changes. A recent overview of different aspects of ocean acidification is the book by Eisler [3]. It should be noted that in comparison to the marked freshwater acidification – which could be seen as more than two-unit decrease in pH, wiped out fish populations in recipient waters of Europe and North America altogether [4–7] (and continues to do so in Asia) and is mainly associated with atmospheric deposition of sulphur and nitrogen oxides – the pH changes predicted for ocean acidification, mainly resulting from changes in carbon dioxide equilibriums, are much smaller.

3.2 What causes ocean acidification The acid-base balance of the marine environment is mainly maintained by a carbon dioxide–carbonic acid–bicarbonate–carbonate balance. The scheme of carbon dioxide-carbonate balance and its disturbances is given in Fig. 3.1.

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0.5 %

89 %

10.5 %

ca. 0 % CO2 1. H2O + CO2 2. H2CO3

2H+ + CO32– H+ + HCO3– 3.

4.

Fig. 3.1: Carbon dioxide balance in the marine environment. (1.) The algae (mostly unicellular) in surface waters consume carbon dioxide in their photosynthesis. This accounts for nearly half of the total global photosynthesis. Any environmental contaminants that decrease the efficiency of this photosynthesis will, thus, increase atmospheric carbon dioxide tension. Atmospheric carbon dioxide is in equilibrium with aquatic carbon dioxide (2.). At the present pH 8.1–8.2, there is approximately 0.5 % carbon dioxide, 89 % bicarbonate and 10.5 % carbonate in marine water. Any increase in carbon dioxide decreases the pH (increases the acidity), as it increases the proportion of the acid component of the equilibria (3.). The proportion of carbonate is affected both by upwelling of carbonate-rich water from close to the bottom sediments as well as calcification and death of calcified organisms, whereby carbonate is removed from the equilibrium (4.).

Several reviews have detailed the historic [8] and present [9] carbon dioxide cycle of the oceans. Both have been reviewed in detail by IPCC [10]. The present carbon cycle contains an anthropogenic component that appears to be decisive for ocean acidification. While some ocean acidification has probably occurred in the past, the present changes in carbon dioxide balance are larger and much more rapid than any effects that have occurred earlier [11]. One significant point to notice is that the location influences the carbon dioxide balance markedly: thus, the overall effects are different in the arctic and tropical oceans with different temperatures [12–14], and coastal areas are affected by both ocean circulation and freshwater inputs [15, 16]. These points are discussed further below.

3.2.1 Effect of atmospheric carbon dioxide loading In simple terms, ocean acidification is mainly caused by an anthropogenic increase in atmospheric carbon dioxide. The pre-industrial carbon dioxide level was approximately 280 ppm and has increased to the present 400 ppm. All predictions agree that the atmospheric carbon dioxide level is continuing to increase, but the level reached by 2100 depends very much on the measures taken to curb emissions, ranging from around 420 to 1000 ppm [17]. The major processes that affect the atmosphere-sea carbon dioxide fluxes are the following. First, the physicochemical dissolution of carbon dioxide into surface water. This component has traditionally been estimated to be de-

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cisive, and has guaranteed that the carbon dioxide level at sea surface is the same as in the above atmosphere. The ocean’s capacity to take up carbon dioxide from the atmosphere decreases with increasing temperature and with increasing carbon dioxide concentration [10]. Traditionally, oceans have been carbon dioxide sinks (e.g. [10]; thus, the anthropogenic effects on temperature and carbon dioxide concentration decrease this “sink” property, when it would be needed most. Second, the circulation of marine water affects the removal of carbon dioxide from the surface water to the deeper parts of the oceans because an increase in temperature decreases water circulation by increasing stratification. Third, the balance of carbon dioxide is affected by its involvement in the functions of biota (e.g. primary production, formation of calcium carbonate shells/skeletons, sinking of calcified structures to sea bottom, etc.).

3.2.2 Influence of primary production An increase in carbon dioxide tension would, in the absence of other changes, tend to increase primary production [18]. Consequently, in the absence of modifying factors, the efficiency of carbon dioxide removal by increased primary production would be enhanced with an increase of carbon dioxide loading. However, because of stratification and reduced nutrient and mineral availability associated with increased temperature, phytoplankton abundance has actually decreased in the recent past [19]. In addition to stratification and decreased mineral availability, primary production is reduced as a result of especially oil and metal pollution [20]. Notably, although tanker wrecks receive much attention, they make up only a few % of the total influx of oil components in oceans [21]. In addition to point sources, contaminants that may affect primary productivity are found everywhere in the world’s oceans [22].

3.2.3 Carbon balance in coastal areas The carbon balance of coastal areas has been reviewed in detail [16, 23–26]. Depending on the location, coastal waters can be either a sink or a source of carbon. The inflow of river water influences the carbon balance of the surrounding coastal waters directly, as many rivers are a rich source of organic carbon [27]. Riparian inputs also affect the total alkalinity of coastal areas, and agricultural practices are decisive in this regard [28]. Total alkalinity affects the carbon dioxide-bicarbomate-carbonate equilibrium, and, consequently, the total amount of carbon (dioxide) that the coastal water contains. Rivers also carry extensive amounts of nutrients, which increases the primary production in coastal waters. The net effect of eutrophication on the carbon dioxide balance depends on the ratio between photosynthesis and respiration. If photosynthesis predominates, the carbon dioxide level decreases. However, often the microbial degradation of organic material is associated with massively increased respiration

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with a net production of carbon dioxide [16]. Another significant effector of carbon balance in coastal areas is water circulation. The upwelling of water from the bottom can make nutrients (and carbonate) available for pelagic organisms. Tidal flats can be both a source and a sink of carbon dioxide. Overall, the different factors affecting the carbon dioxide-carbonate equilibria in coastal zones are highly dynamic, whereby at one period of time conditions favor coastal acidification but at another any general acidification is prevented [16]. Differences can occur on a daily or seasonal basis.

3.2.4 Interactions between temperature changes and ocean acidification Although an increase in atmospheric carbon dioxide level per se would increase primary production, and, other things staying constant, a temperature increase would do the same, the overall primary production in the oceans decreases with increasing temperature [29, 30]. Consequently, temperature increase decreases the effectiveness of the ocean in converting carbon dioxide to oxygen. Although coastal waters constitute only a small proportion of total oceanic area, shelf seas are very important carbon dioxide sinks [31] (but as discussed above, can also have a net acidification effect). An increase in temperature increases the thermal stratification of water, and the fluxes of minerals and nutrients that the primary producers require are reduced in the surface waters [2, 32, 33]. The increased stratification also directly decreases the absorption of carbon dioxide from the atmosphere by reducing its transport from the surface to deeper parts of the ocean.

3.3 Processes of animals that are expected to be affected 3.3.1 pH regulation As discussed above, ocean acidification occurs due to disturbances of the carbon dioxide equilibrium. The same chemical reactions between carbon dioxide and water also occur in the extracellular and intracellular level of animals. In fact, carbon dioxide is the major acidic end product of animal respiration, and carbon dioxide equilibria are the primary components of the regulation of acid-base balance of animals in the short term. Because of this, animals are vulnerable to ocean acidification, and effects of an elevated carbon dioxide level on organismal pH have commonly been demonstrated (e.g. [34, 35]). Organismal pH needs to be tightly regulated, as the three-dimensional structure of virtually all proteins and thereby their function is sensitive to the charge (i.e. protonation) of amino acids, which itself depends on pH [36–38]. Most of the research used to explain effects of ocean acidification on the acid-base balance consists of acute studies. The changes in carbon dioxide and pH (hypercapnia) have been rapidly imposed and typically last only for a few days, and the effects

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on acid-base equilibria and associated ion balance are followed [39]. However, actual ocean acidification is long-term, with noticeable changes of pH only recognizable after years have passed. This difference is important in terms of the regulatory mechanisms involved in controlling acid-base balance. Whereas in short-term hypercapnia experiments any pH disturbance is first buffered and then existing proteins carry out the functions aiming to adjust the acid-base balance (where carbon dioxide equilibria are mainly affected), the acid-base responses to true ocean acidification involve changes in gene expression and even changes in the genetic makeup of the exposed populations. Also, whereas in the short term, acid-base responses that involve either saving the energy (metabolic depression) or increased oxygen consumption may be useful for tolerating external acidification, they cannot be sustained when animals need to reproduce or encounter other environmental stresses. Consequently, most of the existing research on acid-base disturbances (and indeed most aspects of physiology) as a result of increased carbon dioxide level needs to be used with caution when it is implemented to interpret questions pertinent to ocean acidification. The different processes involved in regulating acid-base balance of animals are given in Fig. 3.2 (see also [40] and [41]. Most of the acid loads of animals are caused by the production of carbon dioxide in respiration (0.7–1 molecules of carbon dioxide produced for every molecule of oxygen consumed); it is excreted as carbon dioxide to the surrounding water via the gills (for a detailed review of carbon dioxide excretion see [42]). Gills are also the major site of acid-base relevant ion transporters. The first step that can be regulated is the metabolic production of acid (carbon dioxide) loads. Reducing them via metabolic depression would decrease the acid-base disturbances that need to be corrected in the face of acidification (and increased carbon dioxide levels). Several animals respond to changes in the physicochemistry of water by lowering their metabolic rate [43–46]. Also, it would be expected that animals with lower metabolic rates would be less affected by ocean acidification than animals with higher metabolic rates. Interestingly, this may suggest that polar species are more vulnerable than other species, if the concept of metabolic compensation holds [47, 48]. In such a case, polar species would produce more acidic end products at a given temperature, and consequently would need to spend more energy in pH regulation than other species in an acidified ocean. To our knowledge such a possibility has not been experimentally explored. Also, it is possible that genetic differentiation among the populations of a given species would be targeted to this particular trait, which would be important for the adaptation of a species to ocean acidification. Again, we are not aware that this possibility has been explored. The acid produced in the metabolism then needs to be buffered. In this context, it should be remembered that buffering will not prevent pH changes, only minimize them. While the studies of animals have traditionally concentrated on bicarbonatecarbon dioxide buffering, other types of buffers, e.g., phosphates, free amino acids, dipeptides, peptides and proteins, can also be important. For example, haemocyanin and hemoglobin are important buffers. Typically, the buffering capacities of different

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1. CO2

CO2

H+

Buffer– 2.

H+

Na+

BufferH H+ 4. 3.

K+

Na+

H+ K+ Na+

–

Cl–

Na H+

HCO3–

Cl–

Buffer–

+

HCO3

Cl–

–

HCO3

Fig. 3.2: The principles of treating acid products in an animal. Carbon dioxide constitutes the most abundant acidic end product as it is produced in aerobic respiration in mitochondria. Its excretion occurs via free diffusion through the gills (1.) as discussed in detail by Tufts and Perry [42]. Protons, which can be metabolically produced or result from general acidification, are first buffered by intracellular buffering (2.). If the acid loads cannot be extruded from the cells, the buffering capacity will be depleted, and the intracellular pH decreases more and more. Protons will be extruded from the cells via membrane transport. The most important proton extrusion pathway appears to be the secondarily active sodium/proton exchanger. The secondarily active transporter (depicted in yellow) uses an actively maintained ion gradient. In the case of sodium/proton exchange, an electrochemical gradient for sodium, generated by the sodium/potassium ATPase (primarily active transporter which uses ATP to transport ions against an electrochemical gradient, depicted in red), is used to transport protons against their electrochemical gradient. Equivalent to proton extrusion is bicarbonate influx. It can occur via the passive chloride/bicarbonate exchange (depicted in blue). Bicarbonate transport can also be secondarily active. Transporters that couple bicarbonate movements to sodium transporter have been described. The acid loads are transported from tissues to excretory sites in circulation (where they are buffered). The major excretory sites in water-breathing animals are the gills, where both apical (water-facing) and basolateral (blood-facing) transporters are involved in acid excretion. In addition to the transporters involved in the general intracellular pH regulation, primarily active proton pumps can play a role in acid excretion.

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tissues vary and buffering capacities of aquatic animals are lower than those of terrestrial ones [41]. In addition, the buffering capacities of invertebrates are lower than those of vertebrates, e.g., haemocyanin in invertebrates has a much lower buffering value than hemoglobin in vertebrates. Further, the levels of non-bicarbonate buffers in the haemolymph of invertebrates are generally lower than in vertebrates [43, 46]. The buffering capacity is largest at pH values close to the pK of the buffer molecule. This means that the carbon dioxide-bicarbonate system is a poor buffer in a closed system at physiological pH values (7–8), since its apparent pK value is 6.1–6.3 (depending on the temperature). Thus, virtually all carbon dioxide is in the form of bicarbonate, and acid loads cause little change in the proportion of acid and base forms of the system. The effectiveness of bicarbonate-carbon dioxide stems from the fact that the system is open in the respiratory surface, whereby the volatile carbon dioxide produced can effectively be removed from the animal (see e.g. [41]). The extracellular accumulation of HCO−3 will increase the effectiveness of carbon dioxide-bicarbonate buffering, and is consequently one of the first means to defend against acidification in both invertebrates and vertebrates [43, 46, 49–52]. However, the bicarbonate concentration of extracellular space does not exceed 50 mM [53]. It has been suggested that higher levels than this generate an energetically unfavorable situation for ion regulation that the animals cannot tolerate in the long run (i.e. the energy expenditure for ion regulation is higher than can be allocated with available energy production) [46]. Since ocean acidification is caused by an increase in the carbon dioxide tension of water, it decreases the effectiveness of carbon dioxide-bicarbonate buffering, as buffering depends on the effectiveness of carbon dioxide diffusion to the environment – dictated by the partial pressure gradient from the animal to the environment. With any buffering system, the changes of pH get larger with depletion of the buffer (i.e. buffering done), so the maintenance of effective buffering requires that the acid can be removed from the cell/tissue/animal. Of the potentially acidifying molecules within the cells, carbon dioxide is removed from the cells, tissues and the animal by diffusion (for a detailed review of carbon dioxide excretion see: e.g. [42]). Since ocean acidification is caused by an increase in the carbon dioxide tension of water, its diffusion speed will be reduced because of the decrease in its partial pressure gradient from the animal to the environment. In addition to carbon dioxide, another gas that has acid-base effects and is excreted by aquatic animals is ammonia. However, its associations to ocean acidification are presently thought to be limited. The excretion of protons (or influx of bicarbonate, which is functionally the same) from the cells or from the animals occurs mainly via ion transporters on the cell membranes. The transporters regulating acid-base balance and ion balance are either the same or functionally coupled. Consequently, environmental disturbances in one will also have effects on the other. The major ion transporters involved in the intracellular pH regulation are Na+ /H+ exchange and Na+ independent and dependent Cl− /HCO−3 exchangers [37]. In addition, primary proton pumps are important constituents of pro-

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ton extrusion mechanisms in gills [54–56]. All the sodium-dependent transport mechanisms use the sodium gradient generated by the primarily active Na+ /K+ ATPase. Thus, they are called secondarily active transporters. Only the sodium-independent Cl− /HCO−3 is thought to be a passive exchanger, but also suggestions for ATP-requiring Cl− pumps have been made [57]. Regardless of the ion transporter utilized in pH regulation, energy is consumed in overall proton extrusion (see Fig. 3.2). The energy cost can be up to 40 % of estimated total energy use (3–40 %) [46]. Consequently, any additional acid loads, such as the ones caused by ocean acidification, will increase energy consumption by animals. This can be an important reason for the decreased growth and increased mortality that has been shown to occur in mollusks, arthropods, chaetognatha, echinoderms, sipunculids and vertebrates exposed to increased carbon dioxide and reduced pH (for a review see [51]). Similarly, energy available for protein synthesis may be reduced because of ocean acidification [43]. The effect of ocean acidification on the energy costs of pH regulation becomes aggravated whenever the animals’ activity causes the need for increased proton excretion. Consistent with this, it has been noted that swimming capacity and escape performance is reduced due to ocean acidification [46, 58]. In view of the above, the maximal capacity for pH regulation (which will be related to the availability of/possibility to allocate energy for pH regulation) is important in terms of animals tolerating ocean acidification. Developing animals, which have poorly developed acid-base regulation, are commonly more vulnerable to ocean acidification than adults. The larvae of, e.g. the green sea urchin Strongylocentrotus droebachiensis [59] and the Pacific oyster Crassostrea gigas [60] have approximately the same extracellular pH as that of the environment. However, sea urchin larvae regulate their intracellular pH using the same bicarbonate and sodium/proton transport systems as adult animals [59]. Generally, invertebrates have lower metabolic rates than vertebrates, and can thus use less energy for pH regulation than vertebrates [43, 46, 51, 61], whereby they are expected to be more vulnerable to ocean acidification. Usually species that live in environments where they regularly encounter changes in carbon dioxide level and pH are able to tolerate acid-base disturbances better than other species. They have, e.g. elevated initial HCO−3 levels [62, 63]. However, these animals are adapted to short term changes in pH and carbon dioxide, which is different from the long-term and sustained elevation of carbon dioxide tension and decrease in pH that takes place in ocean acidification. As an example, the adult purple-tipped sea urchin Psammechinus miliaris lives in tidal pools with daily changes in water pH and carbon dioxide tension, but cannot tolerate even a modest hypercapnia if the carbon dioxide increase is long-term and sustained [64]. Acclimation of pH regulation to ocean acidification may occur via a change in pH regulatory pathways, as this may affect the energy usage of pH regulation. For example, Sipunculus nudus changes from sodium/proton exchange to sodium-dependent chloride/bicarbonate exchange in acidified environments in intracellular pH regulation [46, 65]. Acclimation to ocean acidification can also be epigenetic. If adults of ei-

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ther the green sea urchin, Sipunculus droebachiensis, or orange clownfish, Amphiprion percula, and anemone fish, Amphiprion melanopus, are acclimated to hypercapnia, their offspring grow better in hypercapnic environment than offspring of adults that have not been exposed to hypercapnia [66–69].

3.3.2 Calcification The effects that have attracted most attention with regard to ocean acidification are effects on calcification. This is largely because the most noticeable effects of climate change and ocean acidification concern corals [70]. The exoskeletons and shells of invertebrates are largely comprised of calcium carbonate, so its adequate uptake is necessary for the skeletons’ buildup. This is clearly seen in reduced coral calcification during the recent past [70]. The skeletons are formed from the calcium carbonate minerals of aragonite or calcite. Calcium carbonate as calcite is used in the formation of skeletons/shells of some sponges, brachiopods, echinoderms, some serpulids, most bryozoan and some bivalves (e.g. oysters). Most mollusks, both cold- and warm-water corals and some serpulids have mainly aragonite in their shells/skeletons. Of the major forms of calcium carbonate, calcite appears to be more stable (and less soluble) than aragonite [71]. The calcite and aragonite saturation of water is sensitive to pH, pressure and temperature. The saturation with regard to calcite is higher than to aragonite, but for both types of calcium carbonate, the saturation should be adequate for shell/skeleton formation in tropical/subtropical surface waters (Ω > 1) [72]. However, in cold oceans full saturation may not be reached even at sea surface [72], and with increasing depth the saturation of water with carbonate minerals decreases [13, 16]. Also, the effect of freshwater inflow and the nutrients and other minerals it contains will affect carbonate mineral saturation in coastal shelves, occasionally producing especially aragonite undersaturation [16]. Indeed, in cold waters clear examples of problems in shell/skeleton formation have been observed [13]. Because aragonite saturation is lower than that of calcite, and decreases markedly with decreasing water temperature and increasing pressure, especially the calcification of deep-dwelling coldwater corals has been thought to suffer from ocean acidification [73, 74]. However, the calcification of many cold-water corals (e.g. the Mediterranean ones) appears to be little affected by manipulation of carbon dioxide levels and pH to values expected to occur by 2100 [75–77]. The reason why calcification of certain corals can occur even when the environment should be corrosive for (i.e. cause the dissolution of) [73] the carbon skeletons of the corals is depicted in Fig. 3.3. Calcification is possible in an overall aragonite undersaturated environment, if the pH of the site, where calcification occurs, is high enough [78]; the increase of pH will increase the carbonate saturation (such that calcification is possible) or if calcium (carbonate) uptake is concentrative. Interestingly, the overall mechanism of calcification in corals is, to our knowledge, relatively poorly understood. Further, calcification of corals harboring symbiotic al-

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gae and species without symbionts is probably different. Cold-water corals are usually deep-water ones below the photic zone, and consequently with no symbionts. Although the rate of calcification is a direct function of aragonite saturation (Ω), in corals with symbiotic algae it appears that their photosynthesis is involved in the calcification process, whereby calcification is affected by light [79, 80]. The involvement of algal symbionts in calcification means that, when the corals have lost the symbionts in bleaching, also calcification is reduced as has been observed [81]. Significantly, although calcification should occur if aragonite saturation is above 1, results indicate that calcification has been slowed down with reduced aragonite saturation even with the present saturation values (2–4) in most marine areas [82]. To us, this indicates that the availability of calcium ions for calcification in the calcification site is limiting the rate, and that the rate is affected by the concentrations of calcium and carbonate even when they are available at saturating concentrations. This means, as in the case of cold-water corals that are able to carry out calcification at aragonite saturation below 1, that the calcification site is not in equilibrium with ambient water. This being the case, it appears that the calcification mechanism in corals has adapted to the environmental aragonite saturation. As a consequence, the calcification of warm-water corals with symbiotic algae may be reduced more with a decrease of aragonite saturation than that of cold-water corals, even though the ambient water in the former case remains supersaturated with aragonite and becomes unsaturated in the latter. The described abnormalities of calcification of shells and exoskeletons concern especially arctic mollusks and crustaceans [13]. The calcification disturbances may result from several alternative possibilities. First, it is possible that the calcification of crustaceans and mollusks is associated more closely to the calcium carbonate saturation of ambient water than that of corals. Second, if calcification is energetically costly, the energy requirements for it may become overwhelming in unfavorable environments characterized by, e.g. a decrease in aragonite/calcite saturation. Some reports have clearly indicated an increase of energy consumption, seen as reduced growth [83]. Third, in addition to the actual calcium carbonate, the formation of exoskeleton requires, e.g. some proteins (for a detailed account of the formation of exoskeleton in crustaceans, see [84]. If the formation of the components required does not proceed optimally, disturbances of overall cuticular structure may be the result. This possibility is closely related to increased energy demand, because energy is required to produce cuticular components.

3.3.3 Development In this section, development is taken as the whole sequence of events from fertilization to adult (often sessile) animal. The effects of ocean acidification on different stages of development are ecologically very important, as they determine, together with gamete output, the recruitment of species. Generally increased carbon dioxide levels

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2. 1.

Fig. 3.3: Hypothetical examples of causing calcification to be apparently independent from the aragonite/calcite saturation of bulk water. 1. The water that is in contact with the calcifying structure (A) has different properties than bulk water. Calcium is actively taken up in this water compartment (red transporters indicate active calcium pumps) and proton/bicarbonate transporters (green transporters) increase the pH of the calcifying water compartment. The increase in pH will increase the proportion of carbonate taking part in calcification. In addition to proton/bicarbonate transport, photosynthesis affects proton balance. The difference between the properties of calcifying and bulk water compartments depends on the activities of the transporters (i.e. possibility for energy use to increase the calcium concentration and pH in calcifying water). 2. The calcifying site faces bulk water, but takes up calcium actively (B). The effectiveness of calcification depends on how much energy can be used for actively increasing the calcium concentration in the calcification site.

and associated acidification is considered to reduce recruitment [85, 86]. However, to encounter the reduced embryonic/larval/juvenile survival generally found, the reproductive output of a coral reef fish, Amphiprion melanopus, increases markedly at elevated carbon dioxide levels [87]. Notably, however, the yolk sacs of the larvae hatching in an elevated carbon dioxide level were smaller than of those hatching at a normal carbon dioxide level [87], suggesting that the fish in the elevated carbon dioxide environment need to start exogenous feeding sooner than fish in a normal carbon dioxide environment. The effects of climate change and ocean acidification on invertebrate development have been reviewed by Byrne [88], but a similar review about fish is not, to our knowledge, available. Regarding the actual fertilization, both no influence [88, 89] and a reduction [85, 90, 91] have been described in invertebrates. The differ-

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ence may indicate a difference between species/groups but can also result from a difference of defining fertilization. It has been reported that sperm motility is decreased by increased acidity, probably because of a decrease in flagellar movements [92–94]. For fish, no effect of pH changes expected for ocean acidification has been observed in actual fertilization [95]. Further, in some cases little effect of expected carbon dioxide elevations on development of fish has been observed [95, 96]. In other studies, fish developing in a high-carbon dioxide environment are characterized by an increased frequency of abnormalities [97, 98], and tissue or organ damage [99, 100]. In addition to these teratogenic effects, many behavioral effects, discussed in more detail in a separate section, have been observed in developing fish [97, 98, 101, 102]. All these effects may result in reduced recruitment [86]. It appears that seasonal effects (on parents) influence the sensitivity of offspring towards an elevated carbon dioxide level [103]. Thus, the seasonality of ocean acidification responses in fish development may be an example of epigenetic regulation. It is the general view that the embryos/larvae of non-calcifying organisms such as fish are affected less by ocean acidification than ones that are already calcified during development. However, this contention does not include the behavioral and other indirect effects that may be decisive for an animal to survive until adulthood. Even if, for example, the extra- and intra-cellular pH in developing organism can be maintained, it is probable that energy consumption for maintenance increases with reduced growth as a consequence [91]. As discussed below, effects on metabolism (oxygen consumption) are an important consequence of ocean acidification, and may play a role in any developmental effect observed. However, as a broad generalization of the data gathered so far, the acidification predicted before 2100 is not likely to cause a direct effect on development apart from corals and possibly mollusks and echinoderms (sea urchins) [3]. Despite this, when combined with additional stressors (changes), effects may occur. The combination of increased temperature and reduced pH has greater effects than either stressor alone, both on development and adult function [104–107]. Developmental effects may thus be important in shaping the distributions of species and their populations in years to come [106].

3.3.4 Oxygen transport and metabolism 3.3.4.1 Oxygen transport The fact that a decrease in pH and an increase in carbon dioxide tension decreases the oxygen affinity of haemoglobin was described more than a 100 years ago [108]. Later, the effect of pH has been shown for virtually all respiratory pigments of almost all animals. Consequently, whenever ocean acidification causes a decrease in blood/haemolymph pH, oxygen loading in gills will be reduced. As a result, the amount of oxygen transported to the tissues will be reduced and the maximal oxygen consumption of the animal will be decreased. However, although this is a potential

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disadvantageous result of ocean acidification, it only occurs if the animals allow the blood/haemolymph pH to decrease. In some animals, a decrease of the pH experienced by respiratory pigments is out of the question. For example, the squids Illex illecebrosus and Loligo pealei have no venous oxygen reserves, and thus any pH-induced decrease of haemocyanin-oxygen affinity would directly cause a reduction in the amount of oxygen carried, and thereby in maximal aerobic metabolism and eventually on survival [109]. The decrease of pH can, however, be prevented by investing more energy in pH regulation. Thus, the ultimate question even in this regard is the possibility of required energy allocation. Although the possibility that ocean acidification as such could affect animal function as a result of changes in the properties of respiratory pigments has been introduced [110], the complex interplay between the properties of respiratory pigments, environmental temperature and environmental oxygen availability prevent making definite conclusions. Typically, as pointed out in Section 3.2, a temperature increase occurs in most cases simultaneously with ocean acidification, and both are normally associated with changes in the oxygenation of water. When these points are further associated with the fact that different animal populations have both different temperature and other geographical ranges and different haemoglobin (respiratory pigment) isoforms [111, 112], teasing apart ocean acidification effects from the other possibilities becomes nearly impossible. Despite this caveat, the results obtained from natural populations suggest that population-level effects may be associated with differences in haemoglobin function. The population differences in haemoglobin function, so far explored mainly in regard to temperature changes [113], may have significant influence on the population responses and consequent genetic adaptation of species with influence on growth and reproduction.

3.3.4.2 Metabolism Some invertebrates are actually able to reduce their metabolism (e.g. cuttlefish, Sepia officinalis [44], Mediterranean mussel, Mytilus galloprovincialis [114], Chilean mussel, Mytilus chilensis [115], common periwinkle, Littorina littorea [116], and the coral, Acropora digitifera [117]. However, as pointed out above, metabolic depression is merely a short-term answer to pH stress [61]. In some animals such as the common limpet Patella vulgata, increased water carbon dioxide tension has no effect on metabolic rate [118]. However, this can either be the result of the facts that the experiments are short term, that hypercapnia causes behavioral alterations or a combination of both. A behavioral change can be possible in the short term, and if indirect effects are not important for the well-being of the animal, but not feasible in the long term or in natural ecosystems with biological interactions. As an example, although a short-term hypercapnia does not affect the oxygen consumption of the jumping conch snail Gibberulus gibberulus gibbosus, its

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escape behavior is influenced [119], suggesting that it would be caught in a natural environment. In most animals, an increase in oxygen consumption is observed whenever the environment is made hypercapnic. This is mainly due to the increased cost of pH regulation, if an animal tries to maintain organismic/intracellular conditions constant. Full compensation of internal conditions seems to be possible at least for fish. Since the increased ion pump activity for pH regulation must be maintained regardless of the activity, the basal metabolic rate increases. Such carbon dioxide-induced increases have been observed in the coral reef fish, Ostorhinchus doederleini and O. cyanosoma, Pacific oyster, Crassostrea gigas, Sydney rock oyster, Saccostrea glomerata, the Antarctic bivalve, Laternula elliptica and blue mussel, Mytilus edulis [60, 96, 120–122]. One particular tissue that needs special consideration when acidification effects are discussed is the heart. This is because any effects on the cardiac metabolism will affect the availability of oxygen in every part of the body [43, 61]. Cardiac output has, indeed, been shown to decrease in hypercapnia in the yellowtail, Seriola quinqueradiata [123]. Also in the intertidal snail, Littorina obtusata and porcelain crab, Petrolisthes cinctipes, which experience circadian changes in pH and carbon dioxide level, experimental hypercapnia (1100–1500 ppm) reduces the heart rate [124, 125].

3.3.4.3 Scope of activity Because the physiological effects of ocean acidification largely boil down to how the required phenomena can be achieved metabolically, a major question is how the overall metabolic scope responds to ocean acidification. Related to this are also the effects on thermal and hypoxia tolerance, since the metabolism of ectothermic animals increases with temperature, and the metabolism of animals becomes oxygen-limited in hypoxic conditions. The effects of ocean acidification on metabolism will cause a reduction of the aerobic scope (the difference between resting and maximal metabolic rate; [126]) of animals. Moreover, the temperature, where the aerobic scope starts to be reduced, decreases [127]. In principle, the resting metabolism of aquatic animals increases exponentially with the temperature; the same happens for the maximum metabolic rate until a certain limit is reached after which the maximal metabolic rate first levels off and then decreases (Fig. 3.4). Since the aerobic scope is the difference between resting and maximum metabolic rate, the aerobic scope basically indicates the energy available for all activities (swimming, reproduction, immune defense, etc.) beyond routine needs. If ocean acidification increases the basal metabolic rate (mainly due to increased energy demand for pH regulation but also calcification) the animal has less energy available for all the other needs, i.e. the aerobic scope is reduced. Also the temperature at which the aerobic scope starts to decline (pejus temperature, Tpej ) decreases, i.e. the thermal niche of an animal becomes narrower (Fig. 3.4). The influence of direct metabolic changes is compounded by any effects on oxygen loading

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Scope for aerobic performance

Tpej

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Fig. 3.4: The effects of ocean acidification on metabolic rate and thermal tolerance. (a) The metabolic rate (MO2 , maximum shown with solid line and resting with hatched line) of aquatic ectothermic animals increases when the temperature of the water increases. (b) By calculating the difference of resting and maximum metabolic rates (from graph (a)), the scope for aerobic performance as well as the thermal window of the animals can be evaluated. In this window, the critical temperature (Tcrit) indicates the temperature where the capacity for aerobic performance is zero and the survival is time-limited; pejus temperature (Tpej) indicates the temperature where aerobic scope starts to decline significantly, and optimum temperature (Topt) is the temperature where the scope for aerobic activity is largest. (c) Ocean acidification has been suggested to increase the resting metabolic rate of animals mainly due to increased energy demand of pH regulation. Therefore, the aerobic scope (d) is reduced, i.e. animals have a decreased amount of energy available for aerobic activities (swimming, reproduction, immune defense, etc.) and also the thermal window is narrower and, e.g. the upper critical and pejus temperatures will decrease [127].

and delivery, as any reduction of oxygen delivery will cause a reduction in the maximal metabolism of cell/tissue/organism [127]. The decrease in the maximal tolerated temperature has been shown with at least Ostorhinchus cyanosoma and the Pacific oyster, Crassostrea gigas [60, 128]. In gilthead seabream, Sparus aurata, it has been shown that hypercapnia can result in available aerobic energy production not being

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enough for the requirements of the fish, so in the short term they must also obtain energy anaerobically [114]. Naturally, this cannot be a sustained solution for the fish, as the incurring oxygen debt must eventually be treated. Again, we must emphasize that the occurrence of several stressors together may cause an effect, although ocean acidification alone would not cause an observable effect. Much of the discussion above has considered increased temperature together with ocean acidification, because these two environmental changes usually occur together. However, interactions between ocean acidification and other stressors such as toxicants also occur. As examples, the toxicity of copper to early life stages of Arenicola (polychaete worm) increases [129] and the cellular stress responses of the bivalves Crassostrea virginica (oyster) and Mercenaria mercenaria increase when animals are exposed to metal while the medium is acidified [130]. While this is partially caused by the effects of acidification in the speciation of metals in seawater [131], it is likely that the simultaneous occurrence of two or more stresses increases the metabolism enough to cause observable disturbances that are not induced by a single stressor.

3.3.5 Behavior Behavioral changes associated with ocean acidification have been reviewed by Briffa et al. [132]. First, complex behaviors are affected by the fact that energy consumption is increased. This will leave less energy for behaviors such as aggression, courting behavior, predator avoidance, etc., which require a great deal of energy [133]. Second, sensing signals leading to behaviors can be disturbed. The simplest case is that pH affects the structure of the sensed molecule so that it cannot be recognized by the receptor. However, often the sensory structures, especially olfactory receptors, are affected [96, 134]. Odors are used in e.g. settlement of coral reef fishes [128], predator avoidance in the juvenile snail Concholepas concholepas, which inhabits tidal areas of rocky shores [135]; sensing them is disturbed by high carbon dioxide level. Also catching prey for some species can be affected by elevated carbon dioxide level/decreased pH, possibly because odor signals from the prey are not sensed properly [101]. Third, decision-making on the basis of sensing can be disturbed by ocean acidification [136, 137]. Especially pathways involving GABAergic neurons appear to be affected [119, 138] influencing, e.g. the anxiety of the fish [139]. While behavioral effects occur both in larvae/juveniles and adults, they are more common and important in early life stages than in adults, especially when considering the recognition of smells of predators and alarm cues [132].

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3.4 Conclusions Ocean acidification is mainly caused by an increase in carbon dioxide entering the marine environment. The pH is estimated to decrease up to 0.5 pH units by 2100. Such a decrease is much smaller than what can occur in acid rain-induced freshwater acidification, where 2-unit decrease of pH may occur. Furthermore, the changes of pH in naturally occurring CO2 -induced acidification are slow – years may be taken before statistically significant pH decreases are observed. These two features of ocean acidification must be taken into account when the relevance of experimental hypercapnia studies is considered. In most studies, the acidification induced is more severe than what is expected to occur. The onset of acidification is usually rapid, and the duration of experiments short (maximally weeks). As a consequence, the responses in the experiments may not be the same as those caused by the natural environmental change. Because of the magnitude of the pH change in ocean acidification, it is likely that phenotypic effects are not observed, but can be prevented by investing more energy to the phenomena potentially disturbed by hypercapnia. Direct effects are only expected for structures in contact with the acidified water, if their properties are not altered during the time course of acidification. Such structures can be the receptors sensing odors, for example. Since the predicted pH changes are slow, it is likely that animals adapt to ocean acidification mainly via alteration of genetic composition of populations. A species can adapt to expected acidification if the genetic makeup of any of its populations either allows for an increase in energy use to maintain constant an internal environment or gives the possibility to change required properties (e.g. the general proton permeability) so that changes in energy use are not required. It is not likely that mutations enabling the species to do the above would accumulate in the few generations available before the animal must adapt to new conditions.

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Turley C, Findlay HS. Chapter 21 – Ocean acidification as an indicator for climate change. In: Letcher TM (ed.). Climate Change. Amsterdam: Elsevier; 2009. p. 367–90. Artioli Y, Blackford JC, Nondal G, Bellerby RGJ, Wakelin SL, Holt JT, Butenschon M, Allen JI. Heterogeneity of impacts of high CO2 on the North Western European Shelf. Biogeosciences 2014;11:601–12. Eisler R. Oceanic Acidification. A Comprehensive Overview. Enfield NH: Science Publishers, St. Helier; 2012. Farmer AM. The effects of lake acidification on aquatic macrophytes – a review. Environ Poll 1990;65:219–40. Gorham E. Scientific understanding of ecosystem acidification – A historical review. Ambio 1989;18:150–4. Reuss JO, Cosby BJ, Wright RF. Chemical processes governing soil and water acidification. Nature 1987;329:27–32.

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4 Alexandrium spp.: genetic and ecological factors influencing saxitoxin production and proliferation 4.1 Introduction Species of the dinoflagellate genus Alexandrium have been studied intensely for more than 30 years, mainly because some species produce the potent marine biotoxin saxitoxin and its analogs (STX). These toxins can accumulate in a wide variety of marine organisms, including mollusks, crustaceans, starfish, octopus, fish, turtles, marine mammals and birds, and have ecosystem-wide impacts (reviewed in [1]). The accumulation of such toxins can lead to the syndrome Paralytic Shellfish Poisoning (PSP) following the consumption of shellfish. In the Philippines, PSP has resulted in approximately 2000 cases and over 100 deaths between 1983 and 1999 [2] due to the closely related species Pyrodinium bahamense. As a result of the high potency of some of the analogs, PSP outbreaks can occur when cell densities are very low, as few as 100–200 cells l−1 [3, 4]. Mitigating this health risk can lead to a serious economic impact on fisheries and aquaculture industries worldwide. For example, in late 2012, a single bloom of Alexandrium tamarense along the east coast of Tasmania resulted in approximately AUD $ 23 million loss to the state fishing and aquaculture industry [5]. Many Alexandrium species have a global distribution and occur in diverse environments ranging from polar seas [6] to brackish tropical and subtropical lagoons [7]. Species of Alexandrium can occur in high densities in each region [8–14]. The ability of Alexandrium species to colonize and proliferate in coastal habitats and shelf regions around the world has been largely attributed to an extraordinary capacity to adapt. Dinoflagellates are, in general, poor competitors in their habitats, for example, exhibiting poor nutrient uptake efficiency and relatively slow growth rates when compared to other phytoplankton such as diatoms [15, 16]. However, Alexandrium species may compensate for this with specific ecophysiological properties. In addition, they perform circadian nutrient-retrieval migrations, exhibit a high prevalence of mixotrophy and produce allelochemicals and toxins targeted against interspecific competitors and predators [17–21]. Each of these factors serves to guarantee their niche within phytoplankton communities and allow them to form seasonal blooms. Particularly, resting cyst production [22], the ability to utilize organic nutrient sources [23] and chemical defence mechanisms [18, 24] allow Alexandrium species to compensate for their competitive disadvantage and form recurrent toxic blooms [25]. Unlike other phytoplankton blooms, Alexandrium proliferations are not always directly related to the dynamics of the abiotic factors primarily driving photosynthesis and growth, such as high inorganic nutrient concentrations, suitable temperatures and light conditions (e.g. [26]). In many cases, Alexandrium bloom formation is tightly linked to hydrographic condi-

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tions selecting for Alexandrium-specific behavioral adaptations, such as motility and vertical migration [27], or promoting physical aggregation of cells [28]. In past decades, research efforts have been focused on understanding the distribution patterns and factors regulating the dynamics of toxic Alexandrium blooms. Many of the local or regional PSP problems are now well characterized (e.g. [29, 30]). Extensive surveys have resulted in forecast models [31] and monitoring strategies [32] that help to reduce the economic losses and assess the ecological risks of toxic outbreaks. The intense research on species of the genus conducted worldwide also revealed that toxic bloom events caused by Alexandrium have increased in the recent past [33], and that new blooms have occurred in previously unaffected areas [34, 35]. Examples of recent expansions of Alexandrium species are the East Siberian coasts [36], the Mediterranean Sea [37] and the Baltic [38]. Human-assisted introductions, particularly of cysts [39, 40], eutrophication [41] and habitat changes [42] have been identified as factors promoting dispersal and bloom formation. Increasingly, the potential effects of climate change on Alexandrium growth, toxicity and bloom formation are also being addressed by scientists (e.g. [43, 44]). Advances in the understanding of the role of climate factors in bloom formation and toxicity will be particularly addressed in this review.

4.2 Alexandrium taxonomy, phylogenetics and species evolution Most species that produce saxitoxins belong to the genus Alexandrium. The type species Alexandrium minutum Halim was described after it caused a water discoloration (a ‘red tide’) in the harbor of Alexandria in Egypt [45]. There are approximately 35 named species, several of which many are very well known and highly studied [8, 25, 46–48]. The discrimination of species of Alexandrium is based on fine details of the thecal plate morphology, as well as cyst morphology and the formation of chains of cells (Fig. 4.1 A, B [46]). Within the genus Alexandrium, most species are morphologically uniform compared to species of other dinoflagellate genera, and small details such as the size and shape of individual thecal plates must be examined to ensure identification [46]. These are outlined in detail in [46]. Species have the plate pattern of an APC (apical pore complex), 4󸀠 , 6󸀠󸀠 , 5󸀠󸀠󸀠 , 2󸀠󸀠󸀠󸀠 , 6C, 9–10S (Fig. 4.1 A, B [46]). A unique combination of morphological characters is present in most species, consisting of: the size and shape of cells; the size and shape of the first apical and sixth precingular plates; thecal ornamentation; the presence or absence of sulcal lists; chain formation; the shape of the apical pore complex; the size and shape of sulcal plates Sp, Sa, and Ssa; and the size and shape of 1󸀠 and 6󸀠 plates. For example, the species A. diversaporum is distinguished based on a unique combination of morphological features: the presence of two size classes of thecal pores on the cell surface; a medium cell size; the size and shape of the 6󸀠󸀠 , 1󸀠 , 2󸀠󸀠󸀠󸀠 and Sp plates; the lack of a ventral pore; a lack of anterior and posterior connecting pores; and a lack of chain

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formation (Fig. 4.1 A, B [49]). As many of these characters are very small and difficult to see, in order to distinguish species, fluorescence microscopy coupled with staining with Calcofluor white, or other techniques for the staining and dissection of thecal plates, are necessary. Scanning electron microscopy can also be used to distinguish species based on their morphology.

5’’

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Alexandrium pacificum (Group IV) Alexandrium australiense (Group V) Alexandrium tamarense (Group III) 100 Alexandrium mediterraneum (Group II) Alexandrium fundyense (Group I) 100 Alexandrium fraterculus 61 Alexandrium tamiyavanichi 99 Alexandrium tropicale Alexandrium affine Alexandrium monilatum 100 Alexandrium satoanum Alexandrium taylori 96 Alexandrium hiranoi 100 92 Alexandrium pseudogoniaulax 96 Alexandrium diversaporum Alexandrium leei Alexandrium margalefi Alexandrium minutum 64 Alexandrium insuetum 62 Alexandrium tamutum Alexandrium andersoni 79 Alexandrium ostenfeldii

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Fig. 4.1: A – Line drawing of the morphology of A. diversaporum, showing general plate pattern typical of Alexandrium, based on Murray et al. 2014. Ventral side. B – Dorsal side of A. diversaporum. C – A phylogeny of species of the genus Alexandrium, based on ribosomal DNA regions LSU, ITS and SSU (unpublished data). The species highlighted in dark grey produce PSTs. The light grey shaded area is the former A. tamarense species complex.

Some morphological characters that have been used in the past to discriminate species of Alexandrium have been found to change over time in culture, or show intraspecific diversity [50]. The position of the anterior attachment pore (aap), also refered to as the connecting pore [46] in the apical pore complex (APC) [50], and the presence or absence of a ventral pore on the first apical plate, had been found to vary within populations of the species that were then known as A. fundyense (Group I), and A. catenella (Group IV) as well as A. minutum, [8, 51–54]. The presence of a connecting pore on the posterior sulcal plate (sp) can vary within species [46, 55, 56]. In some species, such as Alexandrium affine and A. leei, strains with and without large surface pores have been described [11, 57–59]. Several new species have been found over the past 15

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years [49, 60, 60, 61], suggesting that undescribed diversity may still exist within this genus. For this reason, it is becoming increasingly important that morphological descriptions of species of Alexandrium are as thorough as possible, or preferably include information from genetic sequencing of relevant barcoding regions, such as regions of ribosomal (rDNA) regions such as the large subunit (LSU), internal transcribed spacer region (ITS/5.8S) or the small subunit unit (SSU). Taxa identified based on these combinations of morphological features have generally been found to form monophyletic clades in studies based on molecular sequence data, such as regions of the ribosomal DNA array (SSU, ITS1, 5.8S, ITS2, and LSU rDNA, [50, 55, 62, 63]. However, some species have not been supported as distinct taxa based on phylogenies of rDNA regions and may be conspecific with other taxa. For example, A. peruvianum has recently been found to be within the range of both morphological and molecular genetic diversity of the A.ostenfeldii species complex [64], and these authors have suggested that it be considered a synonym of A. ostenfeldii. The species A. cohorticola, A. angustitabulatum, A. hiranoi and A. lusitanicum require reinvestigation, as they are highly genetically and morphologically similar to other Alexandrium species [49, 55, 65]. Intraspecific variability, including several distinct and well-supported clades, has also been found within some species, for example, A.minutum and A. ostenfeldii [63, 66]. A recent study formally revised the Alexandrium tamarense species complex and separated it into five species [48]. Isolates from the A. tamarense species complex were assigned to A. tamarense, A. fundyense or A. catenella based on two main morphological characters: the ability to form chains and the presence/absence of a ventral pore between plates 1󸀠 and 4󸀠 [46]. However, studies have shown that these characters are not consistent and/or distinctive within species, populations and, in some cases, within a single clonal strain. For many years, field and culture studies have been finding cells exhibiting morphologies intermediate between those of A. catenella, A. fundyense and A. tamarense [46, 54, 67–71]. Phylogenies based on regions in the rDNA operon demonstrated that the sequences from morphologically indistinguishable isolates of the A. tamarense species complex partitioned into five well-supported clades [62, 63, 72–74]. These clades were initially named based on their presumed geographic distribution, such as the North American clade, Mediterranean Clade, Temperate Asian clade, Western European clade and Tasmanian clade [62, 73]. The sympatry among representatives from some clades led to a renaming as Groups I–V [63]. Data on morphology, ITS/5.8S genetic distances, ITS2 compensatory base changes, mating incompatibilities, toxicity, the sxtA toxin synthesis gene and rDNA phylogenies were analyzed in a recent study [48]. As all results were consistent with each group representing a distinct cryptic species, they have separated the groups into five species, and assigned species names as follows: Group I, A. fundyense; Group II, A. mediterraneum; Group III, A. tamarense; Group IV, A. pacificum; and Group V, A. australiense [48]. This comprehensive study may serve as a model for a method by which species complexes and cryptic species of

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dinoflagellates can be carefully separated into meaningful individual species. These newly introduced names will be applied in this review. Most phylogenies of Alexandrium have relied on regions of the rDNA array as marker genes [9, 13, 25, 49, 62, 73]. This is due to the fact that other genes, such as mitochondrial genes (cob and cox1) have tended to be very conserved within the genus, and not able to be distinguished at the species level [75]. Few studies have attempted to evaluate the use of other genes for dinoflagellate phylogenetics, including Alexandrium, such as heat shock protein 90 (hsp90) or actin ([76], unpublished data). In general, these genes are more conserved than regions of rRNA in Alexandrium. Some authors have conducted phylogenetic analyses of Alexandrium using concatenated alignments of several genes [55], and these methods have been useful for illuminating with more certainty the relationships amongst clades. The genus Alexandrium has been found to be monophyletic in almost every study yet conducted, although many studies have not used extensive outgroups for comparison. One of the larger clades of Alexandrium in most studies has been highly supported, and consisted of A. satoanum, A. monilatum, A. taylori, A. hiranoi and A. pseudogonyaulax (Fig. 4.1 C [9, 13, 25, 49, 62, 77]). In general, Alexandrium monilatum has appeared to be a sister group to Alexandrium taylori [78]. Alexandrium margalefi was found to diverge the earliest amongst the remaining species [62]. Several species have been found to be part of a clade that has not been well supported, including Alexandrium diversaporum A. leei, A. insuetum, A. minutum, A. tamutum, A. andersoni and A. ostenfeldii [49]. Within this clade, A. diversaporum has been highly supported as the sister group to Alexandrium leei, and both taxa appeared on comparatively long branches [49] (Fig. 4.1 C). Alexandrium leei has frequently been found to show comparatively long branches within the Alexandrium clade [50, 55, 77, 79]. The monophyly of Alexandrium affine/fraterculus/tamiyavanichi clades with the Alexandrium tamarense complex (A. fundyense, A. mediterraneum, A. pacificum, A. tamarense and A. australiense) has been generally fully supported (Fig. 4.1 A) [12, 25, 48, 50, 55, 62, 80]. The five species of the former Alexandrium tamarense species complex are highly supported as a monophyletic group [48], Fig. 4.1 C).

4.3 What are saxitoxins? The production or not of saxitoxins is patchily distributed amongst Alexandrium species (Fig. 4.1, and see below). Saxitoxin and its analogs (collectively known as the STXs, or the paralytic shellfish toxins, PSTs) are amongst the most potent neurotoxins known. STXs selectively block voltage-gated Na+ and Ca2+ channels and are K+ channel gating modifiers in excitable cells, affecting neural impulse generation [1]. The regulatory limits for the safe consumption of shellfish are 80 μg STX equivalents (STXeq)/100 g tissue in most countries [1]. There are more than 50 known analogs of STX [1, 81], each differing slightly from the parent in three categories: the carbamate

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compounds, which include saxitoxin (STX), neo-saxitoxin (NEO) and gonyautoxins 1–4 (GTX); the N-sulfocarbamoyl compounds, which include the C and B toxins; and the decarbamoyl compounds (dcSTX). These compounds vary in toxic potency by orders of magnitude due to certain minor structural differences. Some of them, such as GTX1, can be at least as potent as STX [82]. It is produced through a series of at least ten reactions involving enzymes with the catalytic activities of a class II aminotransferase, amidinotransferase, SAM-dependent methyltransferase, hydroxylases and O-carbamoyltransferase [83, 84]. In vitro biosynthesis studies confirmed that arginine, acetyl-CoA, SAM and carbamoylphosphate are precursors for STX biosynthesis and led to a revised proposed pathway [85, 86]. Following synthesis of the parent molecule STX, additional enzymes then synthesize the different analogs.

4.3.1 Which species produce saxitoxins? The species of Alexandrium that are well known to produce STXs are Alexandrium minutum, A. ostenfeldii, A. pacificum, A. fundyense, A. australiense, and A. tamiyavanichi (Fig. 4.1 C). These species are not all within the same clade of Alexandrium (Fig. 4.1 C) and do not even represent all species of the former A.tamarense species complex. The patchy distribution of STX production amongst Alexandrium species complicates the study of the evolution and ecology of toxin production. For four species of Alexandrium, A. affine, A. andersonii, A. taylori and A .leei, it is not yet clear whether they have the ability to produce STXs. Reports of strains capable of STX production have generally not been replicated by other studies. One study has reported a strain of A. affine that can produce low levels of STX [59], while other studies have found that strains do not produce STXs [58, 87–89]. It has been reported that a strain of Alexandrium andersonii can produce low levels of STXs [90], while other studies based on both this same strain and other strains using established STX detection methods have not detected STXs [9, 55, 88, 91]. One study reported strains of Alexandrium leei that produced very low levels of STX [59], although this species has been previously found to be non-toxic [11]. Some species are now established to include both clear STX-producing strains and those that do not produce STXs, which sometimes partition into separate clades in phylogenic analyses. These species are Alexandrium minutum [66, 92], Alexandrium ostenfeldii [93] and Alexandrium australiense (previously A. tamarense Group V) [56]. In the species A. pacificum, A. fundyense and A. tamiyavanichi, all strains investigated to date appear to produce STX and analogs (i.e. [8, 25, 94]), with the exception of a sub-clonal strain of A. pacificum, that was the subject of a mating experiment [95]. The quantities of toxin produced and the toxin profiles can differ substantially among strains of the same species for A. pacificum and A. fundyense [8, 25, 48, 56, 96–98]. The profile of Alexandrium minutum appears to be more stable and tends to consist mainly

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of GTX1,4 [98, 99], with minor components of STX and GTX2,3. The most abundant toxin analogs in Alexandrium pacificum (Group IV) from the Mediterranean, Asian, Australian and South American waters tend to be C1,2 and GTX1,4 [48, 98]. Alexandrium fundyense tends to have a similar profile, often consisting of C1,2 and GTX1,4 dominating the other toxins; however, strains can differ substantially, and some do possess high concentrations of STX and NEOstx [8, 25, 48]. In Alexandrium species, decarbamoyl derivatives (dcSTX, dcNEO dcGTX1-4) and the N-21 sulfocarbamoyl analogs C3 and C4 tend to be rarely found [25]. There are several environmental and growth cycle specific factors that appear to influence the amount of STX found in the cell (cellular toxin quota). All previous studies have reported low-level constitutive production of STX during growth in batch cultures. Certain growth phases have been correlated with higher rates of cellular toxin quotas, including the mid-exponential growth phase [25, 99] or the stationary phase [101]. These might be related to the accumulation of toxins throughout cell growth in batch culture. One factor that has been found to lead to an absolute increase in STX production rates is the presence of copepods and their waterborne cues [21, 102–104]. Experiments with copepods have found that their presence can induce STX production in some Alexandrium species, leading to a severalfold increase in production. These studies indicate that STX may play an ecological function as a protection against metazoan grazers [21, 102–104]. While it is known that the cellular toxin quota within a strain may change slightly throughout growth and in relation to culture conditions, or more dramatically, in relation to copepod cues, toxin profile (the proportion of the particular toxin analogs) produced by each species was thought to be relatively stable. However, recent analyses have shown that strains of Alexandrium can change their toxin profile depending on the environmental stress to which they are exposed [105–107]. Therefore, the toxicity of a strain might change, as the profiles shift from more toxic STX to less toxic analogs, or vice versa.

4.3.2 The sxt genes in dinoflagellates Putative genes involved in the biosynthesis of STX, encoded in the dinoflagellate genome, have now been found. This was not simple, as dinoflagellates possess amongst the largest genomes of any eukaryote (approximately 1–200 Gb) and are genetically very different from cyanobacteria, which also possess STX genes; this caused difficulty in attempts to directly sequence genes via PCR-type approaches [88, 93, 108–111]. Speculation had arisen for many years over the production of STXs in dinoflagellates: for example, could they have been produced by co-occurring or intracellular bacteria instead? Studies had found compounds similar to STX analogs in bacteria that lived in dinoflagellate cultures [112]; however, there were doubts in other studies about the identity of these compounds as they have not been confirmed by HPLC and standards, suggesting that they were saxitoxin “imposters” [113–115].

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Additionally, the production of saxitoxin in Alexandrium cultures was found in the absence of bacteria [116]. Genes putatively related to STX biosynthesis from dinoflagellates that have been extensively sequenced were found to have typical dinoflagellate characteristics, such as a conserved 26 bp spliced leader sequence at the 5󸀠 end of the transcribed mRNA (SL ref) and eukaryotic poly A tails [88, 111]. A complement of 14 “core” genes (sxtA-sxtI, sxtP-sxtR, sxtS, and sxtU) is common between the sxt clusters of cyanobacterial STXproducing strains [112], of which eight (sxtA,sxtB, sxtD, sxtG, sxtH or sxtT, sxtI, sxtS, and sxtU) may be directly implicated in STX synthesis [86]. All genes of the cluster putatively necessary for saxitoxin biosynthesis have been found in dinoflagellates, in the species Alexandrium fundyense, A. minutum, Gymnodinium catenatum and Pyrodinium bahamense [88, 110]. sxtA is thought to be the initial gene in the STX synthesis pathway [88]; it is found in all dinoflagellate STX producing organisms, as is the putative second gene in the pathway, sxtG [111]. sxtA has four catalytic domains with different predicted activities: a methyltransferase (sxtA1); N-acetyltransferase (sxtA2); acyl carrier protein (sxtA3); and an amidinotransferase (sxtA4) [86]. The origin of this unique enzyme may be chimeric: the domains sxtA1-3 are most similar to extant proteobacterial sequences, whereas sxtA4 appears to have a separate origin, possibly in actinobacteria [116]. The sxtA gene sequenced in dinoflagellates shows two isoforms – one of which comprises four domains, sxtA1–sxtA4, while the other one encompasses only the domains sxtA1– sxtA3 [88]. The domain sxtA4 appears to be necessary for STX biosynthesis, and therefore, it is possible that only the isoform with this domain is active in catalyzing STX biosynthesis [112]. Part or all of sxtA was found to be absent from other investigated dinoflagellate genera and families that cannot produce STX [48, 49, 88, 111]. Strains of Alexandrium ostenfeldii that do not produce STX appear to lack the domain sxtA4, while those that can produce STX have this domain [93]. sxtG was also found to be present in all STX producing dinoflagellates; however, it was also thought to be present in some closely related non-STX-producing species of Alexandrium [111]. It was absent from other genera of dinoflagellates, apart from Gymnodinium catenatum, which also produces saxitoxin [111]. The domain structures of sxtG appear to be shared between cyanobacteria and dinoflagellates [111]. As convergent evolution of domain architectures is rare, with 0.4–4 % of sequences estimated to be involved in such events, this indicates that convergent evolution is unlikely to have led to this situation. A growing amount of evidence is pointing towards highly expressed genes in dinoflagellates existing in multiple copies, from tens to hundreds (reviewed in [112]. From research to date, it appears that the gene sxtA is no exception, and has been found in the order of 102 copies in species of Alexandrium pacificum, with a relatively similar copy number between strains [108]. This gene has can be detected in field samples using qPCR, and this method facilitates our understanding of Alexandrium proliferation and ecology, and has applications for seafood safety monitoring [108].

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4.4 Ecological factors influencing Alexandrium spp. proliferation and toxicity 4.4.1 The role of ecophysiological adaptations in ecology and bloom formation of Alexandrium life cycles Approximately half of the known and 70 % of the saxitoxin-producing Alexandrium species include benthic-resting cysts in their life cycles [117]. These species regularly alternate between a pelagic growth phase, in which cells divide asexually, and a benthic resting stage, a hypnozygotic cyst. The transitions between these two major stages are typically coupled to switches between reproduction modes and ploidy levels [118]. For many regions experiencing regular Alexandrium blooms, a link to cyst distributions in the sediments has been established [119–122]. Life cycle transformations play a significant role in bloom dynamics of Alexandrium: germination patterns determine the timing of seeding and bloom initiation [123, 124]. Mechanisms regulating cyst formation can in turn influence the magnitude of blooms [125] and even lead to the termination of the growth phase [126]. It has been shown that species or habitat-specific germination and encystment mechanisms can shape bloom strategies of Alexandrium species [127]. Besides determining the distribution and dynamics of blooms, cysts act as dispersal vectors as suggested by molecular analyses revealing introductions of genotypes from distant areas [128, 129]. Resting cysts also contribute to the maintenance of genetic diversity, which has implications for adaptability and resilience of Alexandrium populations. High diversity levels and temporal diversity switches observed in Alexandrium bloom populations have been interpreted in relation to sexual processes involved in cyst formation and germination [96, 130, 131].

4.4.2 Mixotrophic nutrition Although toxic bloom formation by Alexandrium can be promoted by high inorganic nutrient concentrations, particularly in coastal embayments affected by anthropogenic nutrient loading [41, 132], most bloom events are not primarily related to inorganic nitrogen or phosphorus pulses. Productive coastal systems are often characterized by considerable pools of DOM (dissolved organic matter) containing DOC, DON and DOP [133] from allochtonous and autochtonous sources. Alexandrium proliferations have sometimes been observed shortly after the decline of diatom blooms developing on inorganic nutrient pulses [134]. With support from experimental studies, this has been interpreted as the result of increased dissolved organic nutrients released from the decomposing phytoplankton biomass [135]. DON has been shown to increase growth of several Alexandrium species [136, 137]. Evidence for the ability to utilize organic nutrients also comes from enzyme activity assays. Urease activity increases in A. fundyense under conditions of N starvation, indicating that urea is uti-

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lized as an alternative N source [138]. Similarly, utilization of organic phosphorus has been inferred from alkaline phosphatase activity of Alexandrium cells in a phosphorus deficient natural phytoplankton community [139]. Besides N and P, Alexandrium can also take up complex carbohydrates and utilize their C alternatively or complementary to photosynthesis [23]. Phototrophic Alexandrium species have been found to be even capable of phagotrophy [140]. Food vacuoles have been observed in cells of A. ostenfeldii from field samples collected during low production periods [141].

4.4.3 Allelopathy Alexandrium species are known to produce a variety of bioactive compounds. They produce cellular saxitoxins, spirolides or gymnodimines which may play a role in deterring grazers, as has been shown for STX and copepods [21, 102, 104]. Alexandrium species also excrete lytic extracellular secondary metabolites that are effective against a wide range of organisms [18, 24, 142]. These substances can negatively affect and even suppress growth of microbial communities [143], phytoplankton [142, 144] and co-occurring heterotrophic protists [18, 24]. Allelopathy is widespread among Alexandrium species [18]. Allelopathic interactions of Alexandrium with other phytoplankton typically lead to temporary elimination of competitors by triggering, immobilization or cyst formation [145], but they can also be fatal for the target species [18, 24, 146]. Interestingly, within Alexandrium populations, allelopathic potency can vary significantly among individual clones. Within A. tamarense and A. ostenfeldii populations, some clones may not be lytic at all, while others have lethal effects at very low cell concentrations [97, 142, 147]. Allelochemical compounds may also function as grazer deterrents and may have substancial positive effect at the intra-population level as they mutually facilitate conspecifics within Alexandrium populations and therefore may support the population growth and bloom formation in patches of higher cell densities [146]. Exudates of A. ostenfeldii have been shown to alter the swimming behavior of a tintinnid ciliate [52]. This species also induces behavioral disturbance and incapacitation in mesozooplankton and thus deters grazers [148]. For a long time, a possible role of the actual toxins produced by Alexandrium, particularly STXs and spirolides, was investigated in allelochemical interactions. However, evidence for such a mechanism could never be found [18, 24]. The ability to actively deter potential competitors and grazers has been considered an essential aspect of the Alexandrium niche, promoting bloom formation [17, 18, 25, 146].

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4.5 Effects of environmental factors on Alexandrium proliferation and toxicity 4.5.1 Nutrients Blooms of toxic phytoplankton and Alexandrium, particularly, are increasing in their global distribution [4, 25, 149]. It has been suggested that this might be partially driven by high anthropogenic loadings of N and P into the water [4, 24, 149]. PSP toxins (PSTs) are nitrogen-rich alkaloids and their production has been shown to rely on the relative availability of nitrogen. The PST cell quota increase upon phosphorus limitation and decrease upon nitrogen limitation [44, 103, 106]. In a meta-study, it was shown that cellular N:P ratios depend upon phosphorus limitation or nitrogen limitation and the toxin cell quota such as PST content in Alexandrium [44]. Under relative nitrogen limitation, the cellular quota of PSTs decreased, while upon relative P-limitation the cell quota of PSTs increased. Therefore, the total toxin concentrations in natural waters might increase with eutrophication depending on the limiting nutrient as low N:P ratios promotes the internal accumulation of PSTs [44].

4.5.2 Temperature Species of the genus Alexandrium have a wide temperature distribution and represent different types of thermal adaptation. Alexandrium has been reported from all climatic zones [6, 11], and toxic events associated with Alexandrium blooms occur at high as well as at low latitudes. Species with a very wide distribution, such as A. minutum or A. ostenfeldii, are typically comprised of different ecotypes or regional populations with distinct temperature adaptations that reflect the temperature conditions of their bloom niches [150]. Some Alexandrium species seem to be restricted to either cooler or warmer habitats. Toxic outbreaks of A. fundyense are confined to temperate climates and have not been reported from sub-tropical waters [48], whereas PSP toxicities of A. tamiyavanichi are only known from tropical waters [11, 94]. Although the specific temperature windows for blooms are determined by complex factors and may not exactly reflect optimum temperatures for growth, they are dependent on the physiological temperature range, i.e. the specific thermal reaction norm, of the respective regional population. The importance of population-specific temperature adaptation for the understanding of region-specific responses of phytoplankton to environmental change have been increasingly emphasized [151]. This has also become apparent from the numerous ecophysiological studies investigating the effects of temperature on growth and toxicity of Alexandrium from different geographic regions. Alexandrium ostenfeldii isolates from Danish coastal waters have a narrow temperature window for growth which peaks at 20 °C [152], while in a Chinese isolate of this species growth

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is supported through a wide temperature interval with an optimum at 24 °C [153]. Similarly, a narrow temperature interval at the lower end of the temperature gradient was determined for an isolate of A. fundyense from southern Chile (identified at the time as A. catenella) [154]. Generally, it seems that reaction norms of warm temperate Alexandrium populations are wider, compared to cold temperate populations [150, 155, 156]. Such different thermal acclimation patterns will result in different effects of sea surface warming. Studies on warm water Alexandrium isolates showed that a temperature increase of 4 °C (as predicted for the end of the 21st century by climate change scenarios [157]) from ambient bloom temperatures of approximately 20 °C would promote the growth of warm water adapted A. pacificum and A. ostenfeldii [43, 158]. For cold temperate populations with a narrow thermal reaction norm such as the Chilean A. fundyense [154], the projected temperature increase during the bloom season would exceed the temperature tolerance limits and probably result in a restriction or a seasonal shift of the bloom window. When assessing temperature effects on growth of Alexandrium, it needs to be considered that these may be modified by the interactive effects of several factors. Temperature effects can be enhanced by increased pCO2 [158] or alleviated by suboptimal salinities or nutrient conditions [156, 158]. As a consequence of temperature increases, it is expected that the bloom window, the seasonal period of time in which species can proliferate, of many harmful phytoplankton species will expand temporally [159, 160]. Based on the temperature threshold of 13 °C for bloom formation of A. fundyense, a +4 °C warming would prolong the annual time interval for possible A. fundyense outbreaks in Puget Sound by more than a 100 days [161]. However, the effect of warming on the timing of Alexandrium blooms may not be determined by the physiological temperature ranges alone. In a recent modeling study, climate warming was found to modulate the timing and magnitude of dinoflagellate blooms through complex effects on temperature-dependent life cycle processes with feedback effects on successive blooms [162]. A 3 °C temperature increase would lead to earlier and intensified germination, earlier and enhanced cyst formation and higher cyst deposition in dinoflagellates with temperature-regulated life cycle transformations [162]. The temperature regulation of life cycle processes has also been demonstrated for Alexandrium and considered particularly important in shallow coastal habitats [22]. Moderately increased temperature, for example, shortens the mandatory dormancy interval and enhances germination success of A. fundyense [163], A. tamarense [164] and A. minutum [165]. High temperature favors high frequency germination of A. pacificum; this is a prerequisite for bloom formation in Thau Lagoon [166, 167]. Temperature effects on cyst formation, on the other hand, could restrict bloom intervals in time. Temperature-induced cyst formation has been suggested to be an important trigger of bloom termination in Alexandrium populations from South Japan [155] and the US East coast [168]. Again, warming effects on life cycle transformations may be complex. The stimulating temperature effect on cyst formation of A. minutum is particularly pronounced at low salinities [169].

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The expansion of geographic ranges, thought to be one of the most serious consequences of climate warming for harmful algal blooms, is not yet well documented for Alexandrium species. However, reports of toxic Alexandrium blooms from areas with no prior history of PSP toxicities indicate that Alexandrium blooms are spreading to new habitats and that the spreading could be related to changes in temperature. Recently, A. ostenfeldii started to form recurrent toxic blooms in the Baltic Sea, and in a Dutch creek, at water temperatures > 20 °C [35, 38, 134]. Whether the expansion is a result of a recent spreading to new habitats or reflects promotion of background populations by changing temperature conditions, remains a question. Population genetic analyses of Baltic A. ostenfeldii blooms are in support of the latter hypothesis. They revealed a high level of hierarchic genetic differentiation between local bloom sites, which suggests a long-standing presence of the species in the Baltic Sea facilitating differentiated local adaptation [170]. Increasingly, the relevance of global warming for a potential expansion of HABs in polar waters has been discussed [30]. Little is presently known about the distribution of Alexandrium in Antarctic waters [171]; however, successful isolation and characterization of Alexandrium ‘catenella’ from the Antarctic sea has been reported. In the Arctic, the presence of several Alexandrium species has been documented for the Russian Arctic in the past [6]. More recent reports of toxic isolates from Greenland [141, 172] and the Chukchi Sea, detection of shellfish toxicity in Alaska [30] as well as cyst surveys reporting substantial Alexandrium seed beds from Bering and Chukchi Seas [173] suggest that the Arctic seas might be highly sensitive to PSP outbreaks when temperatures rise above thresholds necessary for bloom development. At high latitudes of the southern hemisphere, Alexandrium species are increasingly being detected and found responsible for shellfish toxicities [154, 174]. Dense blooms of A. tamarense have been observed at the coasts of Kamtchatka and Chukchi peninsulas [175]. However, neither the mechanisms behind bloom formation in cold high latitude waters nor the consequences of associated toxicities are well understood at the moment. Most of the toxic effects of Alexandrium species on the environment will depend on the density and duration of blooms as well as the temperature effects on growth rates and distribution. However, temperature can also have direct effects on toxin production in Alexandrium cells. Generally, toxin cell quotas are high at low temperatures, which has been explained by the uncoupling of cell division and toxin synthesis, leading to higher toxin synthesis than cell division rates at lower temperatures [99]. Most of the recently published temperature manipulation experiments have confirmed an inverse relationship between temperature and cellular saxitoxin concentration [103, 154, 158, 159, 176]. Several studies indicate that the toxin-enhancing effect of lower temperatures is most obvious at temperatures ≤ 10 °C. Relatively stable low total toxin concentrations have been found in A. fundyense at temperatures between 10 and 25 °C in contrast to the high levels measured at 5 °C [176]. In A. ostenfeldii, a temperature increase of 4 °C from 20 °C did not lead to net changes in cellular saxitoxin concentrations [43]. Little variation was found among cellular toxin concentrations of

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A. pacificum at temperatures between 18 and 30 °C [98]. These authors detected particularly low toxicities at the lowest tested temperature (12 °C) and interpreted this as a reflection of the low growth rate encountered at this temperature, suggested by the significant general correlation of growth rate and cellular toxin content identified from extensive experimental data on their strain [98]. Contrasting observations emphasize that a unifying pattern cannot be established for the direct or indirect (growth rate) effects of temperature on the cellular saxitoxin quota [99]. This might be attributed to modifying effects of factor interactions [98], or natural variability in response patterns not covered by single strain experiments. Considerable intraspecific variation has been detected in temperature effects on total cellular toxicities of A. ostenfeldii [144]. Previously changes in relative toxin composition in response to temperature modifications have been regarded as an exception. Recent studies, however, increasingly report significant effects of temperature on toxin composition. Shifts in the relative contribution of STX analogs were observed in A. fundyense [176], A. minutum [98], A. pacificum [158] and A. ostenfeldii [144]. In the latter two studies, temperature changes predicted in climate change scenarios led to a relative increase of highly toxic STX, which has ecological relevance as this may amplify the potential harmful effects of intensified blooms in the warmer ocean of the future.

4.5.3 CO2 Only a few studies have investigated the potential effects of increased pCO2 on Alexandrium growth and toxicity. Since photosynthesis and growth of dinoflagellates have been suggested to be CO2 limited due to the low CO2 affinity of their RUBISCO (RUBISCO II), dinoflagellates are generally expected to benefit from higher CO2 concentrations in the water. However, so far the reported responses for Alexandrium are rather inconsistent and vary among the tested species and strains. The growth of A. pacificum was significantly increased in the presence of higher than present day CO2 concentrations in one experiment [158], but not in another [177]. Growth of A. fundyense (previously reffered to as A. tamarense) was not or negatively affected [178], depending on the tested strain. The observed net changes in the growth of A. ostenfeldii were minor [144]. These authors tested the response of multiple strains and found a significant growth rate increase in two of the eight examined isolates, confirming the presence of intraspecific variability [144]. Some of the above-mentioned inconsistencies might be attributed to such intraspecific response diversity and the lack of a representative range of isolates tested. Even when effects on growth are not obvious, elevated pCO2 levels may modify physiological processes in complex ways. Increased rates of net photosynthesis and decreased dark respiration at high pCO2 in A. tamarense have been found [178]. The study furthermore showed that – contrary to the previous assumption that dinoflagellate generally use the same as inorganic carbon source –

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A. tamarense can utilize both CO2 and HCO3− and that uptake of bicarbonate (HCO3− ) is up-regulated under increased pCO2 . In response to higher pCO2 , the kinetics of C acquisition was also modified towards lower Vmax and lower K1/2 . The authors suggested that this strategy, which is different from other dinoflagellates, provides A. tamarense a competitive advantage at low inorganic C conditions, as they potentially occur during dense blooms [179] and support prolonged growth despite C limitation [178]. Changes in pCO2 levels were also found to influence toxicity of Alexandrium spp. Increased pCO2 led to higher cellular PST concentrations and enhanced the production of highly toxic saxitoxin and neosaxitoxin analogs in A. pacificum [158]. The effect was pronounced at low temperature and nutrient limitation. Similarly, increased saxitoxin proportions relative to less toxic GTX2/3 were observed in the high pCO2 treatment of experiments investigating climate change impacts on A. ostenfeldii [144], though cellular PST content was not altered. Both studies emphasized that such shifts towards highly potent PST analogs may exacerbate the toxic impact of future Alexandrium blooms considerably. However, new results reporting contrasting patterns for A. fundyense (former A. tamarense) indicate that more species and individual strains need to be tested to evaluate the suggested link between future CO2 conditions and impacts of Alexandrium toxicity. This study also reported differences among strains in sensitivity of toxin production along the tested pCO2 gradient. Lower toxicities and shifts in analog composition at elevated pCO2 have been demonstrated to involve regulation of genes associated to secondary metabolite and amino acid metabolism as well as sulfatases, explaining the shift in toxin composition to more sulphated analogs [44].

4.5.4 Salinity Many Alexandrium blooms are found in estuaries or coastal areas that may be influenced by river plumes, causing salinity to be reduced due to the influence of freshwater [180]. Many of the Alexandrium spp. for which information is available on salinity tolerance can in fact grow at a wide range of salinities [93, 107, 176, 181]. Similar to temperature, salinity optima typically reflect the bloom niche of a species in the respective habitat [98]. Alexandrium species from coastal waters of the Mediterranean Sea typically have their optima at > 30 psu [98, 182]. In the brackish Baltic Sea, the only low salinity Alexandrium, A. ostenfeldii, grows best at 10 psu [38, 93]. Salinity can be a strong driver of local adaptation and even result in genetic differentiation. Contrasting reaction norms in oceanic and freshwater-influenced A. ostenfeldii populations [93] are reflected by phylogenetic differentiation patterns [64]. To some extent, patterns in reaction norms to salinity seem to be species related. Generally, A. minutum and A. ostenfeldii (syn. with A. peruvianum) have wider salinity reaction norms than the A. tamarense complex or A. tamiyavanichii, which are less tolerant to low salinities. It

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is not clear whether this reflects or defines differences in the habitat ranges between the species. Nevertheless, many studies mention optimum salinities of 25 to 30 for different Alexandrium species, which is the typical range in most periodically freshwater influenced coastal environments where the genus forms blooms [98]. Of these, many species and populations can tolerate salinities of 10 psu and still grow well at 15, implying that freshening of coastal waters as a predicted consequence of climate change (IPCC 2014) will not be physiologically limiting for most Alexandrium spp. [152, 176]. Though growth rates as such may not be affected by lower salinities, associated morphological and physiological traits such as cell size and pigment content might be influenced significantly [98]. Climate-related salinity changes might affect Alexandrium blooms in brackish low salinity habitats. Predicted salinity reduction (from 6 to 4 psu) might drive the species to the salinity tolerance limits [38] and this may counteract the stimulating effects observed at increased temperature and CO2 . Conversely, salinity effects on the growth of Alexandrium are dependent on other environmental factors. At suboptimal temperatures, the optimum salinity interval becomes narrower in the western Mediterranean A. minutum [98]. Though this concerns temperatures that are lower than at present, such interactive effects may generally be responsible for reduced tolerances to decreasing salinities. So far, little is known about the role of factor interaction in the response of Alexandrium spp. to climate change conditions despite the importance of such information for predictions. Salinity effects on toxin content and composition are similarly complex and contradictory as temperature effects. A number of studies report a link between low salinities and high toxin content for different Alexandrium species [159, 176, 181] for A. ostenfeldii and A. tamarense, suggesting that suboptimal salinity conditions favor saxitoxin production. However, such a relationship could not be substantiated as growth rates (reflecting favorable conditions) are typically not correlated with toxin content [101, 150, 176], suggesting that salinity affects toxicity independently from growth rates. Reports of high cellular toxin quota at high and favorable salinities [93, 176], or the opposite [150], support the concept of specific responses to salinity. Also effects on toxin composition are inconsistent, ranging from no compositional changes detected in Canadian A. fundyense [101] to significant changes in individual analog proportions [93]. These authors also reported complete absence of PSTs from high salinity adapted A. ostenfeldii populations, but concluded that this was not a direct salinity effect but genetically predetermined due to the lack of an essential motif of the sxt gene [93]. Very complex changes in analog composition were observed in A. minutum when exposed to different salinities at different temperatures [150], once more emphasizing that toxin patterns are the result of interactive effects of different environmental parameters. With the available information on the relationship between salinity and PSTs, conclusive predictions on consequences of salinity changes for toxicity and possible impacts of Alexandrium spp. blooms cannot be made.

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Freshening of surface waters could affect Alexandrium blooms and toxic outbreaks by modifying the patterns of water column stratification. Dinoflagellates typically benefit from stratified conditions since these lead to the exclusion of non-motile competitors and allow motile dinoflagellates to actively aggregate. Stratification due to river discharge may be associated with increased supply of humic substances, promoting growth of dinoflagellates [160]. Alexandrium blooms are commonly associated with stratified water [182, 184]. In temperate coastal waters, stratification often occurs at times of increased fresh water run-off due to heavy rainfall or ice melt in the spring. River plumes in estuaries have been demonstrated to play an important role in bloom development and shellfish toxicity of A. fundyense in the St. Lawrence Estuary, Canada. Multiyear data revealed that in this system, seasonal patterns of A. fundyense blooms correlate with periods of strong river flushing and rainfall. Low salinity, reflecting freshwater input and water column stability, was statistically related to an increased probability of A. fundyense cells in the water [185]. The relationship of salinity stratification and bloom formation was found to be a result of both the accumulation due to physical processes and active proliferation in the low salinity layer [184]. Interestingly, resting cyst formation of some Alexandrium species seems to be correlated with low salinities [185], which might be a strategy to maintain the seed beds in the low salinity river plume area and assure persistent re-seeding despite offshore advection of surface layer cells into higher salinity waters. The timing and amount of freshwater runoff will potentially affect coastal circulation patterns and the associated advection of Alexandrium inocula from estuarine seedbeds. In the Gulf of Maine this mechanism contributes to the formation of widely dispersed blooms of A. fundyense [186]. Changed salinity conditions in the source area have been considered a factor in interannual variability of bloom development [187].

4.6 Adaptation to changing climate conditions Despite the extensive efforts to characterize the effects of climate related environmental variables on different Alexandrium species and regional Alexandrium blooms, it remains a challenge to predict the response of these harmful dinoflagellates to climate change and assess potential consequences related to their toxicity. “Case”-specific, often contrasting results and reports on factor interactions at different levels have generated a complex picture that cannot easily be explained. This situation is emphasized by the different assessment outcomes regarding climate change impacts on Alexandrium blooms and PST events [161, 188]. These inconsistencies might also point to crucial gaps in our understanding or insufficiency of approaches. Most of the climate change-related environmental effects on proliferation and toxicity of Alexandrium determined so far reflect immediate short-term responses in simplified systems. The respective laboratory studies were typically performed using single isolates. The few studies that have used more than one strain per species [43,

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105, 106, 176] determined in some cases considerable differences in response patterns among the tested isolates. This is in some ways unsurprising, and emphasizes the need to consider intraspecific variability in studies aiming to understand the effects of climate change on Alexandrium species. In fact, variability in ecologically relevant traits and the genetic basis of such phenotypic variability has long been recognized [189] in phytoplankton, and has recently become the subject of intense research in the genus Alexandrium as well. Growing evidence suggests that considerable genetic diversity exists within Alexandrium species populations [97, 131]. When coupled to diversity of adaptively significant phenotypic traits, such standing genetic diversity provides the basis for selection [190]. Large variability in lytic capacity and cellular toxin content was found recently in a bloom population of A. tamarense from the North Sea [97]; this indicates that genetic diversity reflects diversity of the phenotype. Considerable intraspecific diversity in STX production has previously been found with species of Alexandrium [8]. Studies on A. ostenfeldii revealed a surprizing intraspecific diversity of cyst populations in the response to climate factors, suggesting that the seed pool of Baltic A. ostenfeldii populations contains genotypes that will be favored by future temperature, pCO2 and salinity conditions (Fig. 4.2). Some but not all of the favored genotypes had high cellular toxin contents or high proportions of STX. These results indicate that selection from standing genetic variation is an important mechanism of adaptation to changing conditions. Long-term evolutionary response to climate change scenarios, including the incorporation of new mutations into the gene pool, is also an important mechanism of adaptation. Only two studies have so far addressed long-term adaptation in Alexandrium. In one study of A. minutum, evidence was found of change after 2 years of exposure to greenhouse conditions when performing reciprocal growth experiments with acclimated and non-acclimated sub-clones [191]. Interestingly, the evolution of toxicity followed a pattern of neutral evolution, or random mutation, as indicated by large variations of toxin quota among ten long-term adapted sub-clones [191]. The only other long-term study on A. pacificum emphasized the importance of community effects in long-term adaptations [177]. These authors did not detect a direct effect on growth after 12 months of exposure to high CO2 , despite increased competitive success of adapted A. pacificum in reciprocal community experiments [177]. Studies of the regulation and expression of saxitoxin synthesis genes, for multiple clones of more of the saxitoxin-producing species of Alexandrium, will provide important insights into the processes behind the observed effects of environmental variables on toxin patterns. Finally, numerical models accommodating the complexity of interactive factors and their effects on different organizational levels, including life cycle stages, will enable us to better predict the consequences of climate change for Alexandrium blooms and toxicity in different habitats around the world.

4 Alexandrium spp.: genetic and ecological factors |

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Fig. 4.2: Growth rates and corresponding cellular saxitoxin content (total PST and STX) of eight clonal Alexandrium ostenfeldii isolates grown at present day bloom conditions (a, b) and greenhouse conditions (c, d). Asterisks denote significant changes. Data adapted from Kremp et al. (2012).

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[163] Anderson DM. Effects of temperature conditioning on development and germination of Gonyaulax tamarensis (Dinophyceae) hypnozygotes. J Phycol 1980;16(2):166–72. [164] Nı´Rathaille A, Raine R. Seasonality in the excystment of Alexandrium minutum and Alexandrium tamarense in Irish coastal waters. Harmful Algae 2011;10:629–35. [165] Cosgrove S, Nı´Rathaille A, Raine R. The influence of bloom intensity on the encystment rate and persistence of Alexandrium minutum in Cork Harbor, Ireland. Harmful Algae 2014;31: 114–24. [166] Genovesi B, Laabir M, Masseret E, Collos Y, Vaquer A, Grzebyk D. Dormancy and germination features in resting cysts of Alexandrium tamarense species complex (Dinophyceae) can facilitate bloom formation in a shallow lagoon (Thau, southern France). J Plankton Res 2009;3:1209–24. [167] Laanaia N, Vaquer A, Fiandrino A, Genovesi B, Pastoureaud A, Cecchi P, Collos Y. Wind and temperature controls on Alexandrium blooms (2000–2007) in Thau lagoon (Western Mediterranean). Harmful Algae 2013;28:31–6. [168] Anderson DM, Rengefors K. Community assembly and seasonal succession of marine dinoflagellates in a temperate estuary – the importance of life cycle events and predation. Limnol Oceanogr 2006;51(2):860–73. [169] Figueroa RI, Vázquez JA, Massanet A, Murado MA, Bravo I. Interactive effects of salinity and temperature on planozygote and cyst formation on Alexandrium minutum (Dinophyceae). J Phycol 2011;47:13–24. [170] Tahvanainen P, Alpermann TJ, Figueroa RI, John U, Hakanen P, Nagai S, Blomster J, Kremp A. Patterns of post-glacial genetic differentiation in marginal populations of a marine microalga. PLoS ONE 2012;7(12):e53602. [171] McMinn A, Scott FJ. Dinoflagellates. In: Antarctic marine protists. (Scott FJ, Marchant HJ, eds.), pp. 202–250. Canberra & Hobart; Australian Biological Resources Study, Australian Antarctic Division: 2005. [172] Baggesen C, Moestrup Ø, Daugbjerg N, Krock B, Cembella AD, Madsen S. Molecular phylogeny and toxin profiles of Alexandrium tamarense (Lebour) Balech (Dinophyceae) from the west coast of Greenland. Harmful Algae 2012;19:108–16. [173] Natsuike M, Nagai S, Matsuno K, Saito R, Tsukazaki C, Yamaguchi A, Imai I. Abundance and distribution of toxic Alexandrium tamarense resting cysts in the sediments of the Chukchi Sea and the eastern Bering Sea. Harmful Algae 2013;27:52–9. [174] Almandoz GO, Montoya NG, Hernando MP, Benavides HR, Carignan MP, Ferrario ME. Toxic strains of the Alexandrium ostenfeldii complex in southern South America (Beagle Channel, Argentina). Harmful Algae 2014;37:100–9. [175] Selina MS, Konovalova GV, Morozova TV, Orlova T. Genus Alexandrium Halim, 1960 (Dinophyta) from the Pacific Coast of Russia: Species Composition, Distribution, and Dynamics. Russian Journal of Marine Biology 2006;32(6):321–32. [176] Etheridge SM, Roesler CS. Effects of temperature, irradiance, and salinity on photosynthesis, growth rates, total toxicity, and toxin composition for Alexandrium fundyense isolates. Deep Sea Research Part II: Topical Studies in Oceanography 2005;52(19):2491–2500. [177] Tatters AO, Schnetzer A, Fu FX, Lie AAY, Caron DA, Hutchins DA. Short-versus long-term responses to changing CO2 in a coastal dinoflagellate bloom: implications for interspecific competitive interactions and community structure. Evolution 2013;67:1879–91. [178] Eberlein T, Van de Waal DB, Rost B. Differential effects of ocean acidification on carbon acquisition in two bloom-forming dinoflagellate species. Physiologia Plantarum 2014;151(4): 468–79. [179] Hansen PJ. The effect of high pH on the growth and survival of marine phytoplankton: implications for species succession. Aquat Microb Ecol 2002;28:279–88.

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[180] Therriault JC, Painchaud J, Levasseur M. Factors controlling the occurrence of Protogonyaulax tamarensis and shellfish toxicity in the St. Lawrence Estuary: freshwater runoff and the stability of the water column. In: Anderson DM, White AW, Baden DG, editors. Toxic Dinoflagellates. New York; Elsevier Science: 1985. p. 141–6. [181] Grzebyk D, Béchemin C, Ward C, Vérité C, Codd G, Maestrini SY. Effects of salinity and two coastal waters on the growth and toxin content of the dinoflagellate Alexandrium minutum. J Plankton Res 2003;25:1185–99. [182] Bravo I, Vila M, Maso M, Ramilo I, Figueroa RI. Alexandrium catenella and Alexandrium minutum blooms in the Mediterranean Sea: Toward the identification of ecological niches. Harmful Algae 2008;7:515–22. [183] Bravo I, Fraga S, Figueroa RI, Pazos Y, Massanet A, Ramilo I. Bloom dynamics and life cycle strategies of two toxic dinoflagellates in a coastal upwelling system (NW Iberian Peninsula). Deep-Sea Res Pt. II 2010;57(3–4):222–34. [184] Fauchot J, Levasseur M, Roy S, Gagnon R, Weise AM. Environmental factors controlling Alexandrium tamarense (Dinophyceae) growth rate during a red tide event in the St. Lawrence Estuary (Canada). J Phycol 2005;41:263–72. [185] Weise AM, Levasseur M, Saucier FJ, Senneville S, Bonneau E, Roy S, Sauvé G, Michaud S, Fauchot J. The link between precipitation, river runoff, and blooms of the toxic dinoflagellate Alexandrium tamarense in the St. Lawrence. Can J Fish Aquat Sci 2002;59:464–73. [186] Anderson DM, Stock CA, Keafer BA, Bronzino Nelson A, Thompson B, McGillicuddy DJ, Keller M, Matrai PA, Martin J. Alexandrium fundyense cyst dynamics in the Gulf of Maine. Deep-Sea Res Pt II 2005;2(19–21):2522–42. [187] McGillicuddy Jr DJ, Townsend DW, He R, Keafer BA, Kleindinst JL, Li Y, Manning JP, Mountain DG, Thomas MA, Anderson DM.Suppression of the 2010 Alexandrium fundyense bloom by changes in physical,biological,and chemical properties of the Gulf of Maine. Limnol Oceanogr 2011;56(6):2411–26. [188] Bresnan E, Davidson K, Edwards M, Fernand L, Gowen R, Hall A, Kennington K, McKinney A, Milligan S, Raine R, Silke J. Impacts of climate change on harmful algal blooms (HABs). MCCIP Science Review 2013:236–43. [189] Brand LE. Genetic variability and spatial patterns of genetic differentiation in the reproductive rates of the marine coccolitophores Emiliana huxleyi and Gephyrocapsa oceanic. Limnology and Oceanography 1982;27:236–45. [190] Barrett RDH, Schluter D. Adaptation from standing genetic variation. Trends in Ecology and Evolution 2007;23:38–44. [191] Flores-Moya A, Rouco M, García-Sánchez MJ, García-Balboa C, González R, Costas E, LópezRodas V. Effects of adaptation, chance, and history on the evolution of the toxic dinoflagellate Alexandrium minutum under selection of increased temperature and acidification. Ecol Evol 2012;2:1251–59.

Susanna A. Wood, Jonathan Puddick, Hugo Borges, Daniel R. Dietrich, and David P. Hamilton

5 Potential effects of climate change on cyanobacterial toxin production 5.1 Introduction Cyanobacteria are a group of ancient oxygenic photosynthetic prokaryotic organisms originating between three and four billion years ago. They have been reported in a wide range of environments, including oceans, lakes and rivers, as well as extreme habitats such as geothermal springs, desert soils and the Polar regions [1]. Cyanobacteria can exist as solitary, free-living cells or as colonies/filaments consisting of several to thousands of cells enclosed in mucilage. Whilst most cyanobacterial colonies/ filaments and single cyanobacteria cells are microscopic, large populations become visible as mats, crusts and blooms. Populations may be planktonic (Fig. 5.1 (a, b)), suspended in the water column, or benthic, growing on bottom substrate and sometimes forming extensive mats (Fig. 5.1 (c, d)) Extensive blooms or benthic mats of cyanobacteria can be monocultures or more commonly consist of multiple species. They have been associated with increases in nutrient concentrations (i.e. eutrophication), which has been linked to human activities such as de-forestation, agriculture and urbanization. Blooms may be catalyzed by other factors which have been linked to climate change, including increased water temperature, variations in precipitation, extended droughts and increased carbon dioxide levels (see Chapter 7).

(a)

(c)

(b)

(d)

Fig. 5.1: Images of (a and b) planktonic cyanobacterial shoreline scums (Microcystis sp.) and (c and d) benthic Phormidium mats.

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There are approximately 2000 cyanobacteria species described worldwide and more than 50 are known to have strains which produce natural compounds that are toxic (cyanotoxins). Cyanotoxins are a threat to humans when ingested (via water supplies or accidental swallowing) or from contact (dermal or inhalation). Cyanotoxins exhibit a wide range of toxicity mechanisms including: hepatotoxicity; nephrotoxicity; neurotoxicity; and dermatotoxicity [2]. Cyanotoxins can be divided into three broad groups based on chemical structures: cyclic peptides (microcystins and nodularins), alkaloids (cylindrospermopsins, saxitoxins and anatoxins) and lipopolysaccharides (LPS; Tab. 5.1). Tab. 5.1: Cyanotoxins produced by freshwater cyanobacteria (Adapted from Tab. 3.1 of reference [2]). Cyanotoxin

Primary target in mammals

Structural characteristics

Microcystins

Liver Protein phosphatase inhibitors

Cyclic heptapeptide

Nodularins

Liver Protein phosphatase inhibitors

Cyclic pentapeptide

Cylindrospermopsins

Liver Protein synthesis inhibitors

Tricyclic alkaloid with attached uracil

Saxitoxins

Nerve axon Blockage of voltage-gated sodium channels

Tricyclic alkaloid

Anatoxin-a

Nerve synapse Agonists of nicotinic acetylcholine receptors

Bicyclic secondary amine

Anatoxin-a(S)

Nerve synapse Acetylcholinesterase inhibitors

Organophosphate alkaloid

Lipopolysaccharides

Dermis Irritant of all exposed tissue

Lipopolysaccharide

5.1.1 Microcystins and nodularins Globally, microcystins are the most frequently found cyanotoxins [2]. Microcystins are cyclic peptides (Fig. 5.2 (f)) and to date, more than 100 variants have been isolated and characterized [3]. Microcystins and nodularins both contain the unique beta-amino acid Adda, but whilst microcystins are heptapeptides, nodularins contain only five amino acids (Fig. 5.2 (e)). Many cyanobacteria genera produce microcystins [2]. In contrast, nodularins have only been reported to be produced by Nodularia spumigena, primarily a brackish-water species [2], as well as a Nostoc species that grows symbiotically with cycads [4].

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Each microcystin variant differs with respect to the methyl groups and two amino acids within the cyclic peptide ring. This results in pronounced differences in the toxicity of variants. Microcystins and nodularins are hepatotoxins that inhibit protein phosphatases 1 and 2A in affected organisms [5, 6]. Numerous incidents of animal and human poisoning have been attributed to microcystins and nodularins [7]. One of the most severe cases occurred in Brazil in 1996, when water supply to a hospital was contaminated with microcystins. This resulted in 56 fatalities at a dialysis treatment clinic in the hospital [8].

5.1.2 Cylindrospermopsins Cylindrospermopsin is a tricyclic alkaloid (Fig. 5.2 (c)). Two variants of cylindrospermopsin exist; 7-epicylindrospermopsin, a minor metabolite of the cyanobacterium Aphanizomenon ovalisporum [9] and deoxy-cylindrospermopsin [10], which is now thought to have a similar toxicity to cylindrospermopsin. Cylindrospermopsin is a potent protein synthesis inhibitor and causes extensive damage to the liver and kidney [11, 12]. Falconer and Humpage (2001) have also suggested that cylindrospermopsin may act directly as a tumor initiator [13]. Cylindrospermopsin was implicated in one of the most significant cases of human poisoning from exposure to a cyanobacterial toxin when 148 people were hospitalized in 1979 with symptoms of gastro-enteritis after a local water supply on Palm Island (Australia) was dosed with copper sulphate to control a dense Cylindrospermopsis raciborskii bloom [14–16].

5.1.3 Saxitoxins Saxitoxins are alkaloids (Fig. 5.2 (d)) and are fast-acting neurotoxins that inhibit nerve conduction by blocking sodium channels [17]. Saxitoxins are commonly produced by marine dinoflagellates under the name of paralytic shellfish toxins. Whilst saxitoxins from freshwater cyanobacteria have not been implicated in any cases of human intoxication [2], they have killed animals [18] and were identified in an extensive bloom of Dolichospermum (formally known as Anabaena) circinalis on the Murray Darling River (Australia); this resulted in the death of over 1600 sheep and cattle [19].

5.1.4 Anatoxin-a and homo-anatoxin-a These alkaloid toxins (Fig. 5.2 (a)) are powerful depolarizing neuromuscular blocking agents acting through the nicotinic acetylcholine receptor [20]. They are rapidly absorbed when ingested orally. In sufficiently exposed animals, these alkaloid toxins can induce convulsions, coma, rigors, cyanosis, limb twitching, hypersalivation

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and death. Both have been linked with animal and wildfowl poisonings; however, there have being no sufficiently verified reports of human poisonings/fatalities from anatoxin-a [7]. Anatoxins are produced by both planktonic [2] and benthic species [21].

5.1.5 Anatoxin-a(S) Anatoxin-a(S) (Fig. 5.2 (b)), which is structurally different from anatoxin-a, is around ten-fold more potent using the mouse bioassay [2]. It is a cholinesterase inhibitor that induces hypersalivation, diarrhoea, shaking and nasal mucus discharge in mammals [22, 23]. It is thought to be produced only by Dolichospermum lemmermannii [24] and Dolichospermum flos-aquae [23]. Because it is so rarely detected, very little research has been undertaken on this toxin and it is not discussed further in this chapter.

5.1.6 Lipopolysaccharides (LPS) These are integral components of the cell wall of all gram-negative bacteria, including cyanobacteria. The LPS are found in the outer cell membrane and form complexes with proteins and phospholipids [2]. LPS can elicit irritant and allergenic responses in humans and animals tissue [25, 26]. Although comparatively poorly studied, LPS have been implicated in human health problems associated with exposure to cyanobacteria [7]. Because of the limited information on these toxins, they are not discussed further in this chapter. Recent studies suggest that cyanobacteria will thrive under conditions of global climate change (see Chapter 7 or [27–31]). Cyanobacteria are considered the most important and dominant phototrophs in polar freshwater and terrestrial ecosystems, and recent research has shown that toxin production is more widespread than initially thought [32–35]. Climate change therefore has the potential to have a pronounced and dramatic effect on cyanobacterial composition and toxin production. In this chapter, we investigate how climate change may impact biomass. We also consider how climate change could impact cyanotoxin production through three key mechanisms: change in the concentration of toxin-producing species (including the expansion of species into new environments, e.g. from tropical to temperate regions); shifts in the relative abundance of toxic and non-toxic genotypes of bloom-forming species; and variations in the relative amount of toxin produced at a strain level. We review data from both planktonic and benthic species. We also consider a geographic spread including temperate, tropical and polar cyanobacteria.

5 Potential effects of climate change on cyanobacterial toxin production

O O

O H N

O–

P

O

HO H2N

O

OH

O

N

O

O N+

N

N H

N (a)

NH

(b)

N

HN

NH

OH

OH

(c)

O

O

O

OH O

NH NH

O

O

O

NH CH2 O

H N

HN

H N O HO

OH

HN HN

O N

O

O O

O

OH

OH

HN N

N H NH N H

NH

(d)

HO O

H N N H

N

HN

159

NH2 O H

S

O

|

O

O

HN NH2

(e)

HN

NH2

(f)

Fig. 5.2: Chemical structures of several common freshwater cyanotoxins: (a) anatoxin-a, (b) anatoxin-a(S), (c) cylindrospermopsin, (d) saxitoxin, (e) nodularin-R and (f) microcystin-LR.

5.2 Effects of climate change on common toxin producing species In this section, we review four of the most common freshwater planktonic bloom forming genera: Microcystis, Cylindrospermopsis, Dolichospermum (Fig. 5.3 (a–c)), and Planktothrix. Benthic cyanobacteria have also become increasingly problematic in some countries [36]. We therefore use Phormidium (Fig. 5.3 (d)) as a case study to explore how climate change might affect benthic blooms in river environments.

(a)

(c)

(b)

(d)

Fig. 5.3: Common bloom forming cyanobacteria species: (a) Dolichospermum sp., (b) Cylindrospermopsis raciborskii, (c) Microcystis sp. and (d) Phormidium sp.

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5.2.1 Microcystis Microcystis is a ubiquitous genus that has been identified on all continents except Antarctica. Microcystis sp. are well known as a dominant bloom-forming species and they are commonly associated with the production of microcystins. Microcystis cells lack individual sheaths, but under natural conditions are usually organized in colonies that can consist of many thousands of cells which are surrounded by a common mucilage [37]. Microcystis can regulate its position in the water column though gas vesicle production [38, 39] and negatively buoyant carbohydrate stores [40, 41]. Both toxic and non-toxic genotypes exist, and blooms are usually comprised of both; they are only distinguishable via molecular detection of genes in the microcystin synthetase gene operon [42]. Increased water temperature and density stratification induced by climate change [43] are likely to favor proliferation of Microcystis sp. Optimal temperatures for growth and photosynthesis are greater than 25 °C [44–46]. Intense stratification generally favors buoyancy-regulating cyanobacteria, and may also result in more extensive Microcystis blooms [31]. Davis et al. (2009) provided environmental and experimental data to suggest that predicted climatic changes in concert with eutrophication may promote the growth of toxic, rather than non-toxic Microcystis [47]. They conducted surveys at six study sites in temperate northeast United States of America and found that Microcystis became the dominant phytoplankton species at all sites as water temperature approached its annual maximum. They also showed that elevated temperatures resulted in significantly increased growth rates of toxic Microcystis strains in 83 % of experiments, whereas growth rates of non-toxic strains only increased in 33 %. Further evidence of shifts to toxic strains was provided by Dziallas and Grossart (2011) who incubated toxic and non-toxic strains of M. aeruginosa both with and without (axenic conditions) bacteria in the laboratory at three temperatures (20, 26 and 32 °C) [48]. They observed a shift toward toxic strains at the higher temperature. In contrast, Martins et al. (2011) found selection of non-toxic Microcystis over toxic genotypes under favorable environmental growth conditions (including increased temperature) in a temperate reservoir in Northern Portugal [49]. Recent research suggests that rising atmospheric CO2 levels may also affect the composition of toxic and non-toxic genotypes. Van de Waal et al. (2011) conducted chemostat experiments using mixed cultures of toxic and non-toxic M. aeruginosa [50]. The toxic strain reduced dissolved CO2 to lower concentrations than the nontoxic strain, and became dominant at low CO2 levels. Conversely, the non-toxic strain could grow at lower light levels, and became dominant at high CO2 levels, but only under low light conditions.

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5.2.2 Cylindrospermopsis Cylindrospermopsis raciborskii is a solitary, filamentous diazotrophic freshwater cyanobacterium that was originally isolated from the lakes of Java (Indonesia) [51]. It poses a significant threat to ecological and human health because it can form dense blooms and produces cylindrospermopsins and saxitoxins [52]. Initially, it was believed to be confined to tropical environments, but over the past few decades appears to have expanded its range to many temperate regions across the globe [53–55]. Several researchers have suggested that C. raciborskii will become increasingly dominant due to global warming [52, 53, 56]. Growth rates of C. raciborskii respond strongly to temperature between 20 and 35 °C, with maximum rates around 30 °C [57–59]. In tropical environments, C. raciborskii can bloom year-round, but it only dominates in temperate systems during periods of elevated temperatures (i.e. summer months) [54, 55]. In temperate environments it can form spore-like cells (akinetes) that allow it to overwinter and subsequently germinate under favorable temperatures (22 to 24 °C) [60]. Both toxic and non-toxic C. raciborskii strains exist, and among toxic genotypes the amount of toxin produced per cell can vary markedly [57]. The characterization of the cylindropsermopsin biosynthesis pathway [61] and subsequent development of specific QPCR assays [62] have enabled researchers to demonstrate that toxic and non-toxic genotypes co-occur in field populations [63]. To our knowledge, there are no published studies investigating how variables related to climate change will impact the relative abundance of toxin-producing vs. non-toxin producing genotypes, either in laboratory experiments or in the field. Although there is speculation that C. raciborskii blooms will expand and intensify with climate change, the paucity of knowledge on responses of specific toxic genotypes makes inferences regarding the potential toxicity blooms uncertain.

5.2.3 Dolichospermum Dolichospermum (formally known as Anabaena) is a ubiquitous filamentous diazotrophic genus found worldwide. Species within this genus have been reported to produce a suite of cyanotoxins including: microcystins; anatoxin-a; anatoxin-a(S); cylindrospermopsin and saxitoxin [2]. The two most commonly reported bloomforming species are D. circinalis, D. flos-aquae and D. lemmermannii. Dolichospermum most commonly blooms in bodies of water where there is strong temperature stratification. Both water temperature and stratification are predicted to increase with rising air temperature from climate change [56]. Dolichospermum may be selectively advantaged over other species in stratified systems; it can regulate buoyancy to allow for access to nutrients at depth or to regulate light dosage [64, 65]. It can

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also fix atmospheric nitrogen with heterocytes under conditions of low availability of dissolved inorganic nitrogen [66, 67]. Using molecular techniques, researchers have shown that field populations of Dolichospermum comprise both toxic and non-toxic strains [68]. However, there are no published studies that explore how the relative abundance of toxic and non-toxic genotypes will shift with climate change.

5.2.4 Planktothrix Planktothrix is filamentous genus that is usually planktonic, forms solitary trichomes and has gas vacuoles. Two species, P. agardhii (green coloured due to presence of phycocyanins) and P. rubescens (red coloured due to phycoerythrins), are common in the Northern Hemisphere where they form blooms. Some strains produce microcystins [69]. Planktothrix agardhii prefers lakes with low light availability [70]. In contrast, P. rubescens usually blooms in deep oligotrophic lakes, in the low light conditions of the metalimnion [71]. Dokulil and Teubner (2012) analysed a long-term dataset from Lake Mondsee (Austria) to explore the possible impacts of climate change on P. rubescens [72]. From this analysis, they suggest that P. rubescens will only benefit from climate warming early in the year, during late spring overturn and early summer. They also suggest that longer periods of summer stratification are unlikely to favor increasing biomass. The potential impacts of climate change on P. agardhii are still unclear. Bonilla et al. (2012) assembled a large, global dataset for this species [70] and showed that although it was only observed in temperate and subtropical lakes, it could also bloom at a range of temperatures from < 2 °C [73] to 29 °C [74, 75]. Both Planktothrix species can co-exist in water bodies and several studies have investigated the selective advantage of each under field conditions. Planktothrix rubescens is adapted to low light conditions, whereas P. agardhii is more tolerant of higher light intensities [76]. Oberhaus et al. (2007) demonstrated that the combined effects of temperature and light quality and quantity can impact the relative abundance of both species [77]. Planktothrix rubescens tended to be more competitive at lower temperatures (15 °C) and low intensities of green light, whereas P. agardhii was more competitive at higher temperatures (25 °C) and generally less specialized to light quality. Briand et al. (2008a) investigated the effects of temperature, light intensity and nitrate concentrations on several microcystin-producing and non-microcystin-producing P. agardhii strains in monoculture and competition experiments [78]. They observed an advantage of toxic over non-toxic P. agardhii strains under growth limiting conditions (i.e. low light, low temperature or nitrogen limiting conditions). The reverse trend was observed under non-growth-limiting conditions, suggesting that toxic strains may dominate if water temperature rises. Briand et al. (2008b) tracked changes

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in toxic and non-toxic genotypes in a shallow artificial lake in Paris [79]. They found that the abundance of the toxic genotype appeared to be inversely correlated with changes in the density of P. agardhii cells and to a lesser extent, with the availability of certain nutrients and the abundance of cladocerans. These data demonstrate the complexities of predicting the impact of climate change on natural populations.

5.2.5 Phormidium The filamentous cyanobacterial genus Phormidium is found worldwide in diverse habitats ranging from intertidal marshes to Arctic and Antarctic lakes [80]. Under favorable conditions, some Phormidium species form benthic mats. In rivers in some countries, e.g. France and New Zealand, there has been an apparent increase in the prevalence of benthic Phormidium mats [21, 81–83]. When hydrological and environmental conditions are suitable, Phormidium can form thick cohesive black/brown mats that cover extensive areas of the river substrate (Fig. 5.1 (c, d)). Heath et al. (2011) demonstrated that water temperature and river flow were the two main factors driving the presence of Phormidium mats [84]. They observed a marked increase in mats when river flows decreased below one-half of the yearly average and at water temperature > 14 °C. Using an in-stream habitat assessment method, Heath et al. (2013) showed that decreases in summer minimum low flows exhibited negligible change to available Phormidium habitat [85]. The authors concluded that the frequency of flushing flows (a flow that removes the Phormidium from the river substrate), and not flow alone, was critical in determining the presence of Phormidium. Climate change is predicted to cause shifts in precipitation through: (a) strengthening of existing precipitation patterns, i.e. wet regions become wetter and dry regions become drier; and (b) changing storm tracks, which should move away from the equator and toward the poles as atmospheric circulation changes [86]. In France and New Zealand, rivers in drier areas are likely to have prolonged low-flow periods that would be conducive to promoting Phormidium mats. Field and laboratory studies have demonstrated that Phormidium mats are commonly a mixture of toxic and non-toxic genotypes. Among toxic genotypes, anatoxin concentrations can vary up to 100-fold [87, 88]. Heath et al. (2011) undertook a oneyear study at eight sites on two rivers and observed that anatoxin-a and homoanatoxina occurrence was restricted to where water temperature exceeded 13.4 °C [84], and suggested that the toxin-producing strains may ‘out-compete’ non-toxic Phormidium strains under these conditions. More recent research has shown no correlation between water temperature and toxin production [89]. Further culture and laboratorybased studies are required to explore how temperature and other variables indirectly associated with climate change (e.g. river flow, rainfall) may impact toxin concentrations within the mats.

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5.3 Effects of climate change on toxin regulation Environmental variables regulating cyanotoxin production have been a topic of intense interest over recent decades; however, the research has been strongly biased towards microcystin production. Most studies have used laboratory cultures to demonstrate relatively small (three to four-fold) changes in toxin production in response to changes in a number of different variables (e.g. light, nutrients or temperature). More recently, researchers have used field studies to examine the factors regulating toxin regulation [90, 91]. Here we review research on toxin regulation related to variables that are predicted to be altered by climate change: temperature, light, dissolved carbon dioxide concentration and cell density (as a result of increased severity of cyanobacterial blooms).

5.3.1 Microcystins Although the biosynthetic pathway for microcystins is now known [42, 92], there is still uncertainty surrounding the transcriptional and post-translational regulation of the microcystin synthase. Microcystins were initially thought to be feeding deterrents for predators, and studies showing increased microcystin content in cyanobacteria exposed to predators supported this theory [93, 94]. However, the genes responsible for microcystin synthesis pre-date the eukaryotic lineage [95]. Many early studies focused on the effect of temperature on microcystin production and showed that toxin concentrations were maximal around 20 to 25 °C and decreased at higher and lower temperatures [96–102]. More recent research incorporating gene expression has corroborated these results. For example, Dziallas et al. (2011) showed that microcystin transcripts were significantly lower at 32 °C than at 20 and 26 °C in axenic Microcystis cultures [48]. There are strong indications that light may play a regulatory role in microcystin production. Kaebernick et al. (2000) demonstrated that light quality specifically affects microcystin synthase expression, which is initiated at certain threshold intensities [103]. However, transcriptional changes in genes coding for microcystin production changes due to light did not correlate with observed variations in toxin production. Phelan and Downing (2011) showed a strong correlation between microcystin concentration and growth rate under high-light conditions in Microcystis aeruginosa PCC7806 [104]. However, no change was observed at optimal or low-light conditions, and media composition had no significant effect on the relationship between toxins and survival at high-light conditions. The authors therefore suggested a possible role for microcystin was in protection against photo-oxidation [104]. Recently, researchers have provided evidence to support an intracellular function of microcystin in acclimation of Microcystis to high light and oxidative stress [105]. Sevilla et al. (2012) used quantitative reverse transcriptase PCR to monitor changes in the

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levels of transcripts encoding the microcystin-D-synthetase gene (mcyD) in M. aeruginosa PCC7806 subjected to oxidative agents and under different light intensities [106]. The study identified a link between microcystin synthesis and photosynthesis, and indicated that oxidative stress decreased microcystin synthesis. Using a DNA microarray based on the genome of M. aeruginosa PCC 7806, Straub et al. (2011) studied the dynamics of gene expression during the light/dark cycle [107]. They showed that the biosynthesis of secondary metabolites, including microcystins, occurred mostly during the light period, suggesting that microcystins synthesis is connected with photosystem II. Light intensity has also been shown to alter ratios of microcystin variants. A strain of Planktothrix agardhii which produced [Asp3 ] congeners of both microcystin-LR and microcystin-RR exhibited increased levels of microcystin-LR at elevated irradiance levels [108]. The implication of this finding is that microcystin-LR is an order of magnitude more toxic than microcystin-RR, so the resulting phenotype from increased light intensity was more toxic. Another study using batch culture and a natural population of M. aeruginosa found that at a moderate temperature (20 °C) microcystin content would switch from being predominantly microcystin-LR to a less toxic phenotype that was predominantly microcystin-RR. At a higher temperature (28 °C), this did not occur and the levels of microcystin-LR and microcystin-RR production remained constant [101]. A similar observation was reported by Song et al. (1998) using Microcystis viridis, with higher proportions of microcystin-RR at lower temperatures (15–20 °C) [109]. There is very limited data on the impact of pH on toxin production. Van der Westhuizen and Eloff (1983) suggested Microcystis cells were more toxic at high and low pH values corresponding to slower rates of growth [110]. Jahnichen et al. (2007) studied the impact of inorganic carbon availability on microcystin production and suggested that microcystins may be involved in enhancing the efficiency of the adaptation of the photosynthetic apparatus to fluctuating inorganic carbon concentrations [111]. The potential effects of varying carbon dioxide levels on microcystin production from Microcystis are not well known. Van de Waal et al. (2009) showed that an excess supply of both nitrogen and carbon yielded high cellular nitrogen : carbon (N : C) ratios, accompanied by high cellular contents of total microcystin and the nitrogen-rich variant microcystin-RR [112]. They found comparable patterns in Microcystis-dominated lakes, where the relative microcystin-RR content increased with the seston N : C ratio. Using field-based studies Wood et al. (2011) showed that Microcystis sp. can “switch” microcystin production on and off [90]. They also observed and experimentally induced 20-fold changes in microcystin quotas within a five-hour period and measured up to a 400-fold change in microcystin-E-synthetase gene (mcyE) expression [90, 91]. Both of these studies and a recent laboratory study [113] draw a correlation between microcystin production and cyanobacteria cell densities. This suggests that toxin production per cell may increase if blooms and scum formations intensify as is predicted under future climate change scenarios [56].

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5.3.2 Nodularins Little work has been performed on the environmental parameters that affect nodularin production. Lehtimaki et al. (1997) demonstrated that nodularin production was generally highest under conditions that promoted growth [114]. Intracellular nodularin concentrations increased at higher temperature (25 – 28 °C), higher irradiance (45– 155 μmol photons m−2 s−1 ) and higher phosphate concentrations (200–5500 P μg l−1 ).

5.3.3 Cylindrospermopsins A study exploring the effect of temperature (20–35 °C) on growth and cylindrospermopsin content of C. raciborskii grown in batch culture showed a strong negative correlation between toxin content and temperature. When grown at 35 °C, none of the study strains produced any detectable cylindrospermopsin, but when transferred back to lower temperatures, cylindrospermopsin production was restored [57]. Preußel et al. (2009) highlighted how toxin production can vary among strains of the same species [115]. They investigated the influence of light and temperature on cylindrospermopsin production of two Aphanizomenon flos-aquae strains using semicontinuous cultures. A light gradient from 10–60 μmol photons m−2 s−1 , in combination with temperatures of 16, 20, and 25 °C, was assessed. Cylindrospermopsin concentrations showed a significant decrease with increasing temperature in one strain, whilst there was no clear relationship with temperature in the other. They also noted that cylindrospermopsin production at different light intensities varied at the three temperatures. In both strains, cylindrospermopsin concentrations increased significantly at 20 °C for light intensities from 10 to 60 μmol photons m−2 s−1 , whereas they decreased significantly at 25 °C in the same light gradient. They also suggested that conditions which constitute physiological stress (i.e. low light and temperature) triggered active release of the toxin into the medium. Dyble et al. (2006) investigated the role of light intensity on growth and cylindrospermopsin production in C. raciborskii cultures [116]. Maximum growth rates occurred at 75 μmol photons m−2 s−1 , whereas maximum intracellular and extracellular toxin concentrations occurred in cultures grown under the highest light intensity (140 μmol photons m−2 s−1 ). The authors speculated that the highest intracellular toxin concentrations in the field are likely to occur in actively growing C. raciborskii populations exposed to light intensities of 75–150 μmol photons m−2 s−1 for more than two weeks. Few studies have investigated changes in cylindrospermopsin production in benthic cyanobacteria. Bormans et al. (2014) explored cylindrospermopsin production in Oscillatoria sp. PCC 6506 and demonstrated that the total cylindrospermopsin content was highest during the exponential growth phase at intermediate light levels (10 μmol photons m−2 s−1 ) and during the stationary growth phase under lower and higher light levels.

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5.3.4 Saxitoxins Only a small number of studies have explored changes in saxitoxin synthesis in relation to potential climate change. In a study on C. raciborskii (strain T3), saxitoxin synthesis was up-regulated at a light intensity 100 μmol photons m−2 s−1 and not at 50 and 150 μmol photons m−2 s−1 [117]. This study also showed that saxitoxin production exhibited a circadian rhythm under the three light intensities tested. A study on a different C. raciborskii strain (C10) found contrasting results [118]. In this study, the extracellular and intracellular saxitoxin content of this strain was analyzed at two different temperatures; 19 and 25 °C. At 25 °C an additional variant (dcSTX) was detected, but there was no significant change in toxin concentration. Comparing these data to other studies illustrates how responses to environmental parameters may differ among species. For example, Dias et al. (2002) studied Aphanizomenon sp. (LMECYA 31) and showed enhanced saxitoxin production at a high temperature (28 °C) compared to a lower temperature (22 °C) which was typical of in situ conditions [119].

5.3.5 Anatoxins Rapala et al. (1993) studied Dolichospermum and Aphanizomenon and demonstrated that high temperature decreased the amount of the toxin produced regardless of growth rates [120]. Growth-limiting, low-light conditions and growth-inhibiting, highlight conditions decreased the amount of anatoxin in Dolichospermum cells. In contrast, the highest light flux of the study did not limit the growth or decrease the level of the toxin in the cells of Aphanizomenon. Rapala and Sivonen (1998) monitored anatoxin production in Dolichospermum strains and observed that anatoxin-a production was slightly higher under sub-optimal temperature and light levels [96]. Also, under low light, considerable amounts of extracellular anatoxin-a were detected. Araoz et al. (2005) studied Ocillatoria sp. (PCC6506), a benthic anatoxin-producing cyanobacteria, and noted shifts in the relative production of variants at different temperatures and when different carbon sources were utilized [121]. They observed that synthesis of anatoxin-a occurred at 22 °C when there was no carbon dioxide enrichment of the growth medium, while homanatoxin-a was produced when the medium was at 25 °C and supplemented with 10 mM sodium hydrogen carbonate to maintain carbon dioxide enrichment.

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5.4 Climate change and its effect on cyanobacteria and toxin production in Polar environments In Antarctica, cyanobacteria are considered the most important and dominant phototrophs in freshwater and terrestrial ecosystems [122]. Although less well studied, cyanobacteria are also abundant in the Arctic [123]. In both regions (where there is available water), they commonly form benthic or floating microbial mat communities which can be several centimeters thick (Fig. 5.4). Nostocales and oscillatoriales are usually dominant within these mats, and these taxa have adapted to tolerate harsh conditions, including high salinity, intense ultraviolet radiation and repeated cycles of desiccation/hydration [124, 125]. One of the fastest rates of climate warming has been documented in the Arctic and maritime Antarctic [126], with mean annual temperatures over the Antarctic Peninsula having increased by 0.5 °C per decade [127]. Freshwater habitats may be more severely impacted than terrestrial habitats as the relative increase in water temperature has been reported as two to three-fold higher than the corresponding change in local air temperature [128]. A rise in temperature in ice-free areas is expected to increase the availability of liquid water (from glacial and permafrost melting) [129]. Such changes could dramatically increase water availability, possibly facilitating increased cyanobacterial growth. Wood et al. (2008) found high abundance of “dormant” cyanobacteria in Dry Valley soils [130]. As water availability increases, extensive macroscopic growth could occur within a relatively short time frame. Experiments using cloches (small covers that provided protection from wind and increased temperature and water availability) positioned on arid Dry Valley soils provided evidence to support the suggestion that increased water availability will favor cyanobacteria, and showed rapid growth of moss and cyanobacterial biomass (personal communication with D. D. Wynn-Williams as cited in [129]). Recent reports suggest that there is widespread microcystin production in Antarctic microbial mats including from the McMurdo Ice Shelf, Bratina Island [131, 132], McMurdo Dry Valleys [32] and the Antarctic Peninsula [35]. Although the microcystin producing species have not been definitively identified, Wood et al. (2008) used molecular techniques to provide compelling evidence that Nostoc sp. was the producer in at least some of the microcystin-positive samples [32]. Despite the high abundance of cyanobacteria in the Arctic [133], there is only one report of microcystins in cyanobacterial mats [33] and one report in cyanobacteria from lichens in this region [134]. Kleinteich et al. (2012) demonstrated that microcystin production and species diversity increased in laboratory-cultured Arctic and Antarctic cyanobacterial mats when they were exposed to elevated temperatures (8–16 °C compared to 4 °C) [33]. It was suggested that this could be the result of increased toxin production per cell due to more favorable conditions. These conditions might include higher temperatures favoring the growth of toxin-producing species and allowing them to become dominant

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(a)

(b)

(c)

Fig. 5.4: Antarctic cyanobacterial mats: (a) extensive mats in a meltwater pond in Miers Valley, (b) small pinnacles dominated by Leptolyngbya sp. and (c) a large colony of Nostoc sp.

during the experiment, or higher species diversity/abundance that could intensify inter-species competition at higher temperatures. In the case of the latter hypothesis, increased toxin concentrations could serve as a means for quorum-sensing, intraspecies communication or as allelopathic compounds. In addition to the potential ecosystem effects of increased microcystin production, Kleinteich et al. (2012) warned that changes in cyanobacterial diversity could lead to instability of the mat ecosystems, allowing the establishment of non-polar mesophilic species [33]. Although less well studied, several other cyanotoxins have been detected in the Polar regions. Kleinteich et al. (2014) confirmed the presence of cylindrospermopsin and deoxycylindrospermopsin in several cyanobacterial mats collected around Rothera (Antarctic Peninsula), although the concentrations measured were low compared with benthic species of warmer climatic zones [35]. Saxitoxin was detected for the first time in a cyanobacterial mat collected from northern Baffin Island in the vicinity of Cape Hatt in the Arctic [34]. To date, the identity of the toxin producer/s has not been determined. Anatoxins have not been identified in Polar samples, however this may be due to a lack of systematic surveys for this toxin, as several known producers are widespread in these regions (e.g. Phormidium autumnale). Although no data exist on how climate change will impact the abundance and diversity of toxins produced by cyanobacteria in Polar regions, rates of metabolism, nitrogen fixation and photosynthesis of Antarctic cyanobacteria are thought to be optimal around 15 °C [135]. Kleinteich et al. (2012) theorized that production of secondary metabolites may also be optimal at this temperature [33]. It is therefore plausible that climate change could yield increases in the concentration of these toxins, with as-yet unknown impacts on these fragile ecosystems.

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5.5 Conclusions There is increasing evidence that regional and global climatic change will benefit many species of harmful bloom-forming cyanobacteria (see Chapter 7 and [27–30]). It is predicted that there will be increases in cyanobacteria growth rates, leading to greater longevity and severity of bloom events, and shifts in geographic distributions. Research indicates that many of the key bloom forming genera, i.e. Microcystis, Cylindrospermopsis Dolichospermum, will benefit under predicted climatic change scenarios such as rising temperatures, enhanced water column vertical stratification and shifts in seasonal weather patterns (e.g. droughts). It is extremely challenging to make generalizations about how climate change will impact cyanotoxin production. The studies to date have been undertaken using different conditions (e.g. growth media, batch vs. continuous culture), different methods for determining toxin concentrations (e.g. enzyme-linked immunosorbent assay, liquid chromatography-mass spectrometry) and have utilized different strategies for normalizing data to biomass (e.g. cell concentration, dry weight, optical density). Comparable studies have also shown contrasting responses between species (e.g. microcystinproducing Microcystis and Planktothrix), and even among strains of the same species (e.g. saxitoxin-producing C. raciborskii). We have only reviewed studies that have investigated variables closely associated with climate change. Many other variables may be indirectly related to climate change (i.e. nutrients, metal availability) and have also been shown to impact toxin production. Changes in these variables may have an equal or greater impact on toxin production, or act synergistically with the variables discussed. Studies that have undertaken multi-stressor experiments, for example, investigating the effect on toxin production of temperature at a range of light intensities, have shown complex interactions and relationships, reinforcing how difficult it will be to predict the impacts of climate change in the natural environment. Finally, it should be noted that the majority of studies have been undertaken in the laboratory. Whilst these have provided essential and underpinning knowledge on toxin synthesis, and enabled detailed studies of single stressors, important natural variables and interactions that may regulate toxin production are removed. There is an urgent need for further field-based investigations that can better mimic the complex interactions and multiple stressors that influence toxin production in cyanobacteria.

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Acknowledgment The authors thank the New Zealand Ministry of Business, Innovation and Employment (UOWX0505; Lake Biodiversity Restoration), the Marsden Fund of the Royal Society of New Zealand (12-UOW-087), the Royal Society of New Zealand International Research Staff Exchange Scheme grant (MEAT Agreement 295223) and the Marie Curie International Research Staff Exchange Scheme Fellowship (PIRSES-GA-2011-295223) for funding.

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[115] Preußel K, Wessel G, Fastner J, Chorus I. Response of cylindrospermopsin production and release in Aphanizomenon flos-aquae (Cyanobacteria) to varying light and temperature conditions. Harmful Algae 2009;8:645–50. [116] Dyble J, Tester PA, Litaker RW. Effects of light intensity on cylindrospermopsin production in the cyanobacterial HAB species Cylindrospermopsis raciborskii. African Journal of Marine Science 2006;28:309–12. [117] Carneiro RL, dos Santos MEV, Pacheco ABF, Azevedo SMFdOe. Effects of light intensity and light quality on growth and circadian rhythm of saxitoxins production in Cylindrospermopsis raciborskii (Cyanobacteria). Journal of Plankton Research 2009;31:481–8. [118] Castro D, Vera D, Lagos N, Garcıìa C, Vásquez M. The effect of temperature on growth and production of paralytic shellfish poisoning toxins by the cyanobacterium Cylindrospermopsis raciborskii C10. Toxicon 2004;44:483–9. [119] Dias E, Pereira P, Franca S. Production of paralytic shellfish toxin Aphanizomenon sp. LMECYA 31 (Cyanobacteria). Journal of Phycology 2002;38:705–12. [120] Rapala J, Sivonen K, Luukkainen R, Niemelä S. Anatoxin-a concentration in Anabaena and Aphanizomenon under different environmental conditions and comparison of growth by toxic and non-toxic Anabaena-strains – A laboratory study. Journal of Applied Phycology 1993; 5:581–91. [121] Aráoz R, Nghiêm H-O, Rippka R, Palibroda N, de Marsac NT, Herdman M. Neurotoxins in axenic oscillatorian cyanobacteria: Coexistence of anatoxin-a and homoanatoxin-a determined by ligand-binding assay and GC/MS. Microbiology 2005;151:1263–73. [122] Taton A, Grubisic S, Brambilla E, De Wit R, Wilmotte A. Cyanobacterial diversity in natural and artificial microbial mats of Lake Fryxell (McMurdo Dry Valleys, Antarctica): A morphological and molecular approach. Applied and Environmental Microbiology 2003;69:5157–69. [123] Jungblut AD, Lovejoy C, Vincent WF. Global distribution of cyanobacterial ecotypes in the cold biosphere. The ISME Journal 2009;4:191–202. [124] Vincent WF. Microbial Ecosystems of Antarctica. Cambridge; Cambridge University Press: 1988. [125] Cavacini P. Soil algae from northern Victoria Land (Antarctica). Polar Bioscience 2001;14: 45–60. [126] Vaughan DG, Marshall GJ, Connolley WM, Parkinson C, Mulvaney R, Hodgson DA, King JC, Pudsey CJ, Turner J. Recent rapid regional climate warming on the Antarctic Peninsula. Climatic Change 2003;60:243–74. [127] Turner J, Colwell SR, Marshall GJ, Lachlan-Cope TA, Carleton AM, Jones PD, Lagun V, Reid PA, Iagovkina S. Antarctic climate change during the last 50 years. International Journal of Climatology 2005;25:279–94. [128] Quayle WC, Peck LS, Peat H, Ellis-Evans JC, Harrigan PR. Extreme responses to climate change in Antarctic lakes. Science 2002;295:645. [129] Cowan DA, Ah Tow LA. Endangered Antarctica environments. Annual Review of Microbiology 2004;58:649–90. [130] Wood SA, Rueckert A, Cowan DA, Cary SC. Sources of edaphic cyanobacterial diversity in the Dry Valleys of Eastern Antarctica. The ISME Journal 2008;2:308–20. [131] Hitzfeld BC, Lampert CS, Spaeth N, Mountfort D, Kaspar H, Dietrich DR. Toxin production in cyanobacterial mats from ponds on the McMurdo Ice Shelf, Antarctica. Toxicon 2000; 38:1731–48. [132] Jungblut A-D, Hoeger SJ, Mountfort D, Hitzfeld BC, Dietrich DR, Neilan BA. Characterization of microcystin production in an Antarctic cyanobacterial mat community. Toxicon 2006;47: 271–8.

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6 Harmful marine algal blooms and climate change: progress on a formidable predictive challenge 6.1 Introduction In a strict sense, harmful algal blooms are completely natural phenomena that have occurred throughout recorded history (e.g. Exodus, Captain Vancouver in 1793). Whereas in the past three decades unexpected new algal bloom phenomena have often been attributed to eutrophication [1] or ship ballast water introduction [2], increasingly novel algal bloom episodes are now circumstantially linked to climate change. It is unfortunate that so few long-term records exist of algal blooms at any single locality; ideally we need at least 30 consecutive years. Whether the apparent global increase in harmful algal blooms represents a real increase or not is therefore a question that we will probably not be able to answer conclusively for some time to come. There is no doubt that our growing interest in using coastal waters for aquaculture is leading to a greater awareness of toxic algal species. People responsible for deciding quotas for pollutant loadings of coastal waters, or for managing agriculture and deforestation, should be made aware that one probable outcome of allowing polluting chemicals to seep into the environment will be an increase in harmful algal blooms. In countries that pride themselves on having disease and pollution-free aquaculture, every effort should be made to quarantine sensitive aquaculture areas against the unintentional introduction of non-indigenous harmful algal species. Nor can any aquaculture industry afford not to monitor for an increasing number of harmful algal species in water and for an increasing number of algal toxins in seafood products, or to use increasingly sophisticated analytical techniques such as LC-MS. Last but not least, global climate change is now adding a new level of uncertainty to many seafood safety monitoring programs. Climate on our planet has been constantly changing, over scales of both millions of years (glacial to interglacial periods) and short-term oscillations of tens of years (ENSO, NAO). The Earth’s climate in the distant past has at times been subject to much higher ultraviolet-B (UVB) levels and CO2 concentrations than we are seeing at present. The first photosynthetic cyanobacteria evolved 3.5 billion years ago at CO2 levels 1000 times those of the present, followed by green algae 1000 million years ago (mya) at 500 times the present value and dinoflagellates 330–400 mya at eight times the present value, whereas more recently evolved diatoms and haptophytes operated under comparatively low CO2 environments (two to three times the present value) [3]. During the past 800,000 years, atmospheric CO2 has fluctuated between 180 ppm in glacial and 300 ppm in interglacial periods, but in the past 200 years, this has increased from 280 ppm to 400 ppm at present, with values of 750–1000 ppm

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predicted by 2100. In the past 1000 years, our planet has gone through episodes warmer than at present, such as the medieval warm period AD 550–1300, and colder than now, such as the little ice age AD 1300–1900. Global temperatures in the past 20–30 years have increased significantly with a further rise of 2–4 (6) °C predicted over the next 100 years [4]. From a geological perspective, there is nothing remarkable about the magnitude of climate change we are experiencing now, except that it appears to proceed at a faster pace and starts from a warmer baseline. The current rate of changes in carbonate chemistry referred to as ‘ocean acidification’ (at least 10–100 times faster than the recent glacial transitions) however is unprecedented within the last 65 million years. Because of their short generation times and longevity, many phytoplankton are expected to respond to current climate change with only a very small time lag. They are expected to spread quickly with moving water masses into climatic conditions that match the temperature, salinity, land runoff and turbulence requirements of the species. However, our knowledge of marine microalgae’s potential to adapt is very limited. Collins and Bell [5] grew the freshwater green alga Chlamydomonas over 1000 generations at almost three times the present atmospheric CO2 concentration. The cells acclimated to the change but did not show any genetic mutations that could be described as adaptation. Lohbeck et al. [6] exposed Emiliania huxleyi coccolithophorid cultures founded by single or multiple clones to increased concentrations of CO2 in 500 generation selection experiments. Compared with populations kept at ambient CO2 partial pressure, those selected at increased partial pressure exhibited higher growth rates suggestive of adaptive evolution. Organism’s limits to adaptive capacity exist but remain largely unexplored, but we also should not underestimate that microbial life in the oceans had some 3.5 billion years to evolve, thus representing an enormous genetic diversity and physiological plasticity [7]. While we can expect changes in distribution, performance and genetic diversity of individual species, complete extinction is unlikely.

6.2 Algal bloom range extensions and climate change Temperature defines the geographic distribution of many species and their responses to climate change. Shifting temperature means and extremes alter habitats and cause changes in abundance through local extinctions and latitudinal expansions or shifts. Vulnerability is thought to be greatest in polar organisms due to their narrow temperature ranges and in tropical species living close to upper thermal limits. The dinoflagellate Pyrodinium bahamense is presently confined to tropical, mangrove-fringed coastal waters of the Atlantic and Indo-West Pacific. A survey of cyst fossils (named Polysphaeridium zoharyii) going back to the warmer Eocene 50 million years ago indicates a much wider range of distribution in the past. For example, in the Australasian region at present, the alga is not found farther south than Papua New

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IPWP

Recent plankton

Fossil

Recent cysts

NSW, 100.000 years ago

Fig. 6.1: Global distribution of Pyrodinium bahamense in recent plankton (top left) and much wider distribution in the fossil cyst record (bottom left) (after [2]). Top right: The fossil dinocyst Dapsilidinium pastielsii, which became extinct in the Atlantic during cooling in the early Pleistocene but survived in the warm-water refuge of the Indo-Pacific Warm Pool (IPWP) (after [11]). Bottom right: Pyrodinium bahamense cyst from Port Moresby Harbor, Papua New Guinea.

Guinea but, some 100 000 years ago in the Pleistocene, the alga ranged as far south as Sydney Harbour [8]. There is concern that, with increased greenhouse warming of the oceans, this toxin-producing species may one day return to Australian waters (Fig. 6.1). Pyrodinium blooms in 1972 in Papua New Guinea coincided with the fatal food poisoning of three children in a seaside village, diagnosed as PSP (Paralytic Shellfish Poisoning). Since then, toxic blooms have spread to Brunei and Sabah (1976), the central (1983) and northern Philippines (1987) and Indonesia (North Mollucas). There is strong circumstantial evidence of a coincidence between Pyrodinium blooms and unusual weather linked to the El Niño Southern Oscillation (ENSO). In the Philippines alone, Pyrodinium has now been responsible for more than 2000 human illnesses and 100 deaths resulting from the consumption of contaminated shellfish as well as sardines and anchovies [9, 10]. Comparable examples of spreading of dinoflagellate cyst species distributions with increasing temperatures, and shrinkage of biogeographical zones with decreasing temperatures, are known. The fossil dinocyst Dapsilidinium pastielsii became extinct in the Atlantic during cooling in the early Pleistocene, but a warm-water refuge for this taxon was recently discovered in the Indo-Pacific Warm Pool (IPWP: Japan, Indonesia, Vietnam, Palau, Philippines) [11]. Ciguatera caused by the benthic dinoflagellate species complex Gambierdiscus toxicus is a tropical fish-food poisoning syndrome well-known in coral reef areas in the Caribbean, Australia and especially French Polynesia (Fig. 6.2). Whereas, in a strict sense, this is a completely natural phenomenon; from being a rare disease two centuries ago, ciguatera has now reached epidemic proportions in French Polynesia. From

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Atlantic Ocean

Pacific Ocean

Indian Ocean

Fig. 6.2: Global distribution of ciguatera fish food poisoning (from polynesie.demeyer.net). Inset: the causative benthic dinoflagellate genus Gambierdiscus.

1960 to 1984, more than 24 000 patients were reported from this area, which is more than six times the average for the Pacific as a whole. Evidence is accumulating that reef disturbance by hurricanes, military and tourist developments as well as coral bleaching (linked to global warming) and perhaps future increasing coral damage and changing macrophyte cover due to ocean acidification are increasing the risk of ciguatera [12]. In the Australian region, Gambierdiscus is well-known from the tropical Great Barrier Reef and southwards down to just north of Brisbane, but in the past five years this species has exhibited an apparent range extension into South-East Australian sea grass beds as far south as Merimbula (37° S), aided by a strengthening of the East Australian Current [13]. A similar apparent range expanse of Gambierdiscus has been reported in the Mediterranean and the Canary Islands [14], and the Caribbean and West Indies [15]. In the North Sea an analogous shift of warm-water phyto- and zooplankton to the North Pole has occurred due to regional climate warming [16–18].

6.3 Range extensions further aided by ship ballast water transport Ballast water is seawater which has been pumped into a ship’s hold or dedicated ballast tanks to steady it by making it heavier and thus less likely to roll; the water is released when a ship enters port. Ballast water on cargo vessels was first suggested as a means of dispersing marine plankton 100 years ago [19]. However, it was only in the 1980s that the problem sparked considerable interest, after evidence was brought forward that non-indigenous toxic species such as the PST dinoflagellate

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Gymnodinium catenatum had been introduced into Australian waters in sensitive aquaculture areas, with disastrous consequences for commercial shellfish farms [20]. To prove that a particular species of microalga has been introduced into a particular location is much more complex than for example in the case of macroalgae or marine invertebrates. However, the dogma of widespread cosmopolitanism of marine microalgae is now increasingly rejected in favor of underestimated microalgal diversity. The implication is that human-mediated translocations and their impacts may have been seriously underestimated. For example, we now recognize that the Alexandrium tamarense/catenella species complex is comprised of six to eight different genotypes [21, 22], some of which are always or mostly non-toxic (Tasmanian and European clades), while others (North American and Temperate Asian clade) contain potent toxic strains. The molecular detection in the Mediterranean ports of Sete and Barcelona of Alexandrium catenella with a temperate Asian ribotype not found anywhere else in Europe [23, 24] thus most likely reflects a human-assisted introduction. Similarly, Australian populations of Alexandrium hide introduced temperate Asian ribotypes among indigenous nontoxic strains [25]. However, a recent outbreak of Paralytic Shellfish Poisoning in Tasmania (SE Australia) in 2012, which cost the local economy $ 23 M through contamination of shellfish, abalone and lobster, may have been caused by a climate-driven shift in genetic structure of the dinoflagellate populations [26]. It is by no means clear whether the harmful dinoflagellate Cochlodinium polykrikoides, which caused major fish kills and problems for the desalination industry in the Arabian Gulf in 2008, is a ballast water introduction [27]. It could also represent a climate-driven range expansion from 2005 Malaysian blooms, or be a species that has always been there in low concentration but now stimulated by changing environmental conditions such as anthropogenic nutrient enrichment (Fig. 6.3). A concerted effort looking for Cochlodinium cysts or their genetic fingerprints in dated sediment depth cores may help resolve this important question. A comparable scenario applies to the red-tide dinoflagellate Noctiluca scintillans in Australian waters. While native to the Sydney region (since 1860), in the 1980s eutrophication caused it to expand its distribution; subsequently, climate-driven range extension moved it southwards into the Tasmanian region, and only very recently (2008) have we seen it expand against prevailing current systems to Port Esperance and Cairns, hence implying an additional role of domestic ballast water dispersal [28] (Fig. 6.4). Ecosystems disturbed by pollution or climate change are more prone to ballast water invasions than mature stable ecosystems [29].

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Cochlodinium polykrikoides

1995

1978

2004

1999 2000

2008 2004 ?

East Asian Phillipines American Malaysian Unknown

Fig. 6.3: Geographical distribution of the fish-killing dinoflagellate Cochlodinium polykrikoides with years of first bloom occurrences noted. The 2008–2009 Gulf bloom was a new bloom phenomenon for the Gulf region, represented by the American/Malaysian ribo-type, which is distinct from the East Asian and Philippines ribotypes (after [54]). We cannot yet resolve whether this Gulf bloom event represents a ballast water introduction, a climate-driven range expansion or a response of a previously cryptic species to anthropogenic nutrient enrichment, or perhaps a combination of all of the above.

1860– 1950

10–S 1980– 1993

1994– 2005

20–S 2008-13 30–S 40–S 50–S 60–S 100–E

120–E

140–E

160–E

160–E

Fig. 6.4: Australia-wide distribution of the red-tide dinoflagellate Noctiluca scintillans (inset), in the periods 1860–1950 (only known from Sydney Harbour), 1980–1993 (expanding along New South Wales coast in response to eutrophication), 1994–2005 (East Australian Current driven range extension) and 2008–2013 (expanding against prevailing current systems to Port Esperance and Cairns, implying a role for domestic ballast water dispersal. Updated after [28].

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6.4 The formidable challenge of predicting phytoplankton community responses While we made considerable progress in our understanding of the physics of climate change, our understanding of the impacts on biological communities is in its infancy. There will be winners and losers from climate change, but predicting how individual species will respond poses a formidable challenge [30, 31]. Increasing temperature, enhanced surface stratification, alteration of ocean currents, intensification or weakening of local nutrient upwelling, stimulation of photosynthesis by elevated CO2 , ocean acidification and increased frequency of heavy precipitation and storm events causing changes in land runoff and micronutrient availability may all produce contradictory species or even strain-specific responses. Complex factor interactions exist [32] and ecophysiological experiments rarely take into account genetic strain diversity and physiological plasticity [33]. An undoubted key driver for future phytoplankton changes will be increasing sea surface temperature and enhanced water column stratification (shallowing of the mixed layer). In open ocean environments, this may lead to more rapid depletion of surface nutrients, a decrease in replenishment from deep nutrient rich waters and therefore reduced phytoplankton biomass (“oligotrophication” [34]). By contrast, in high-latitude regions with relatively deep mixing and sufficient nutrients, decreasing mixing depth can result in higher phytoplankton biomass because of increased light availability [35]. Cyanobacteria can dominate both marine and freshwater ecosystems under higher temperature, notably when combined with eutrophication [36]. Extreme weather events such as heavy rainfall (nutrients from land runoff), hurricanes and dust storms are well known as impacts to marine phytoplankton. Winds influence the supply of iron to the surface ocean through aeolian transport of dust from land to sea. This contributes micronutrients such as iron, which stimulate Karenia brevis blooms off Florida [37]. Hurricanes have been claimed as being responsible for expanding the distribution of cyst-producing toxic dinoflagellates (e.g. Alexandrium tamarense in New England [38]). In Hiroshima Bay, blooms of the fish-killing raphidophyte Chattonella marina followed typhoon-induced accretion of nutrient-rich land runoff [39]. It is widely predicted that increasing CO2 will lead to ocean acidification (“the other CO2 problem”), a decrease in ocean pH (from 8.1 down to 7.7 by 2100) and associated changes in carbonate chemistry. Most of the tested harmful algal bloom species lack carbon-concentrating mechanisms and hence they may benefit from increased atmospheric CO2 [3]. While initial attention focused on potentially adverse impacts of ocean acidification on calcifying organisms such as the coccolithophorid Emiliania huxleyi [40], a much greater impact may derive from how ocean acidification will alter the availability of micronutrients such as iron [41]. However, nobody could have predicted that ocean acidification would induce more subtle changes such as the swim-

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ming behavior of the fish-killing raphidophyte Heterosigma akashiwo [42]. Dominance and community structure of harmful bloom dinoflagellates can be profoundly altered by changing pCO2 [43], and both toxic dinoflagellates (Alexandrium catenella, Karlodinium veneficum) and diatoms (Pseudo-nitzschia multiseries) have been shown to produce higher cellular toxin concentrations under near-future levels of ocean acidification [44–46].

6.5 We can learn from the fossil record, long-term plankton records and decadal scale climate events The ecosystem response to natural climate variability in the past provides a glimpse into the climate-induced changes of the near future. We can learn important lessons from the dinoflagellate cyst fossil record [47] and from the few long-term data sets available (such as the Continuous Plankton Recorder surveys [16]; Fig. 6.5, Fig. 6.6). Data from the Continuous Plankton Recorder in the Northeast Atlantic confirm that warming from 1960 to 1995 enhanced phytoplankton growth [48]. In response to transient warming, phytoplankton distribution in the North Atlantic shifted towards the pole by hundreds of kilometers per decade since the 1950s. Phenology of plankton in the North Atlantic was also affected, with differences in sensitivity between groups. Hinder et al. [49] attributed a recent decline in North Sea dinoflagellates relative to diatoms to warming, increased summer windiness and thus water column turbulence. Seasonal timing of phytoplankton blooms is now occurring up to four to five weeks earlier in the North Sea in relation to regional climate warming (Fig. 6.6). Similarly, Moore et al. [50] predict longer-lasting Alexandrium catenella blooms in Puget Sound under future climate change scenarios. Coccolithophore blooms of E. huxleyi in the Bering Sea were reported for the first time during the period 1997–2000, probably in response to a 4 °C warming, combined with a shallower mixed layer depth, higher light levels and low zooplankton grazing [51]. A similar range expansion and increase abundance since the 1990s of E. huxleyi into the Southern Ocean has been reported [52], but the underlying mechanism remains to be fully explained. However, not all trophic levels are responding to the same extent, and where zooplankton or fish grazers are differentially impacted by ocean warming, this may have cascading impacts on the structure of marine food webs [53].

6.6 Mitigation of the likely impact on seafood safety The greatest problems for human society will be caused by being unprepared for significant range extensions of HAB species or an increase of algal biotoxin problems in areas that are poorly monitored at present. While, for example, ciguatera contamina-

6 Harmful marine algal blooms and climate change

Anomaly

Anomaly

–0.7

–0.3

Anomaly

0.1

0.5

| 189

Anomaly

0.9

Fig. 6.5: Decadal anomaly maps (difference between long-term 1960–1989 mean and the 1990– 2002 period) for four common HAB species (from left to right) Prorocentrum, Ceratium furca, Dinophysis, Noctiluca (insets) in the North Atlantic. Note the increase in Prorocentrum, Ceratium furca and Dinophysis along the Norwegian coast and increase in Noctiluca in the Southern North Sea (after [17]).

1950

1960

Years (1948–2001) 1970 1980

1990

2000

2

Month

4 6 8 10 12 0.0 0.8 1.3 1.5 1.7 1.9 2.1 2.4 2.8 3.3 Phytoplankton colour index

Fig. 6.6: Long-term monthly values of “phytoplankton color” in the central North Sea from 1948 to 2001. Circles denote > 2 SD above the long-term monthly mean (after [48], with permission). Note an apparent shift towards earlier spring and autumn phytoplankton blooms.

tion would be expected and monitored for in tropical coral reef fish, with the apparent range extension of the causative benthic dinoflagellate into warm-temperate sea grass beds of South-Eastern Australia, other coastal fisheries unexpectedly could be at risk. Similarly, incidences of increased surface stratification in estuaries, heavy precipitation or extreme storm events are all warning signs that call for increased vigilance of

190 | Gustaaf M.Hallegraeff

monitoring seafood products for algal biotoxins even in areas not currently known to be at risk. Only with improved global ocean observation systems, such as improved and expanded ocean sensor capabilities (e.g. argo floats, ocean gliders, coastal moorings and coastal radar, multi-wavelength and variable fluorometers, optical sensors) in support of integrated satellite-derived “ocean color” maps and expanded biological and bio-geochemical observations (continuous plankton recorder, eco-genomics) can we expect to define management options, forecast ocean-related risks to human health and safety, and shed light on the impact of climate variability on marine life and humans in general. It is pleasing to see that a number of national (e.g. the US NSTC Joint Subcommittee on Ocean Science and Technology Ocean Observatories Initiative (OOI), the Australian Integrated Marine Observing System (IMOS)) and international programs (e.g. the Intergovernmental Oceanographic Commission of UNESCO’s GEOHAB, IOC Global HAB status reports initiative) are actively pursuing these ambitious goals.

Acknowledgment Our understanding of climate-driven impacts on phytoplankton processes is continuously increasing. This chapter represents a partial update on an earlier review in J. Phycology 2010, and benefitted from my attendance at an IOC-UNESCO GEOHAB meeting in Paris in April 2013.

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[25] Bolch CJS, de Salas MF. A review of the molecular evidence for ballast water introduction of the toxic dinoflagellates Gymnodinium catenatum and the Alexandrium tamarensis complex to Australasia. Harmful Algae 2007;6:465–85. [26] Campbell A, Hudson D, McLeod C, Nicholls C, Pointon A. Tactical Research Fund: Review of the 2012 paralytic shellfish toxin event in Tasmania associated with the dinoflagellate alga, Alexandrium tamarense. FRDC Project 2012/060 ISBN: 978-0-646-90570-9 [27] Richlen ML, Morton SL, Jamali EA, Rajan A, Anderson DM. The catastrophic 2008-2009 red tide in the Arabian Gulf re.gion, with observations on the identification and phylogeny of the fish-killing dinoflagellate Cochlodinium polykrikoides. Harmful Algae 2010;9:163–72. [28] McLeod DJ, Hallegraeff GM, Hosie GW, Richardson AJ. Climate-driven range expansion of the red-tide dinoflagellate Noctiluca scintillans into the Southern Ocean. J Plankton Res 2012; 34:332–7. [29] Stachowicz JJ, Terwin JR, Whitlatch RB, Osman RW. Linking climate change and biological invasions: ocean warming facilitates non-indigenous species invasion. Proc Nat Acad Sc USA 2002;99:15497–15500. [30] Hallegraeff GM. Ocean climate change, phytoplankton community responses, and harmful algal blooms: a formidable predictive challenge. Journal of Phycology 2010;46:220–35. [31] Glibert PM, Allen JI, Artioli Y, Beusen A, Bouwman L, Harle J, Holmes R, Holt J. Vulnerability of coastal ecosystems to changes in harmful algal bloom distribution in response to climate change: projections based on model analysis. Global Change Biology 2014; doi: 10.1111/ gcb.12662 [32] Feng Y, Warner ME, Zhang Y, Sun J, Fu FX, Rose, JM, Hutchins DA. Interactive effects of increased pCO2 , temperature and irradiance on the marine coccolithophore Emiliania huxleyi (Prymnesiophyceae). European J Phycol 2008;43:87–98. [33] Fabry VJ. Marine calcifiers in a high CO2 world. Science 2008;320:1020–2. [34] Behrenfeld MJ, O’Malley RT, Siegel DA, McClain CR, Sarmiento JL, Feldman GC, Milligan AJ, Falkowski PG., Letelier RM, Boss ES. Climate-driven trends in contemporary ocean productivity. Nature 2006;444:752–5. [35] Doney SC. Plankton in a warmer world. Nature 2006;444:695–6. [36] O’Neil, Davis TW, Burford MA, Gobler CJ. The rise of harmful cyanobacteria blooms: The potential roles of eutrophication and climate change. Harmful Algae 2012;14:313–34. [37] Walsh JJ, Steidinger KA. Saharan dust and Florida red tides: The cyanophyte connection . J Geophys Research 2001;106:11597–11612. [38] Anderson, DM. Bloom dynamics of toxic Alexandrium species in the northeastern US. Limnol Oceanogr 1997;42:1009–22. [39] Kimura T, Mizokami A, Hashimoto T. The red tide that caused severe damage to the fishery resources in Hiroshima Bay: Outline of its occurrence and environmental conditions. Bull Plankton Soc Japan 1973;19:82–96. [40] Riebesell U, Zondervan I, Rost B, Tortell PD, Zeebe RE, Morel FMM. Reduced calcification of marine plankton in response to increased atmospheric CO2. Nature 2000;407:364–367. [41] Shi, D, Xu Y, Hopkinson BM, Morel FM. Effect of ocean acidification on iron availability to marine phytoplankton. Science 2010;327:676–9. [42] Kim H, Spivack AJ, Menden-Deuer S. pH alters the swimming behaviors of the raphidophyte Heterosigma akashiwo: Implications for bloom formation in an acidified ocean. Harmful Algae 2013;26:1–11. [43] Fu FX, Tatters, AO, Hutchins DA. Global change and the future of harmful algal blooms in the ocean. Mar Ecol Prog Ser 2012;470:207–33. [44] Fu FX, Place AR, Garcia NS, Hutchins DA. CO2 and phosphate availability control the toxicity of the harmful bloom dinoflagellate Karlodinium veneficum. Aquat Microb Ecol 2010;59:55–65.

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[45] Sun J, Hutchins DA, Feng Y, Seubert EL, Caron DA, Fu FX. Effects of changing pCO2 and phosphate availability on domoic acid production and physiology of the marine harmful bloom diatom Pseudo-nitzschia multiseries. Limnol Oceanogr. 2011;56:829–40. [46] Tatters AO, Flewelling LJ, Fu F, Granholm AA, Hutchins DA. High CO2 promotes the production of paralytic shellfish poisoning toxins by Alexandrium catenella from Southern California waters. Harmful Algae 2013;30:37–43. [47] Dale B. The sedimentary record of dinoflagellate cysts: looking back into the future of phytoplankton blooms. Sc Mar 2001;65:257–72. [48] Edwards M. Phytoplankton blooms in the North Atlantic: results from the Continuous Plankton Recorder survey 2001/2002. Harmful Algae News 2004;25:1–3. [49] Hinder SL, Hays GC, Edwards M, Roberts EC, Walne AW, Gravenor MB. Changes in marine dinoflagellate and diatom abundance under climate change. Nature Climate Change 2012;2: 271–5. [50] Moore SK, Mantua JM, Hickey B, Trainer VL. Recent trends in paralytic shellfish toxins in Puget Sound, relationship to climate, and capacity for prediction of toxic events. Harmful Algae 2008;8:463–77. [51] Merico A, Tyrrell T, Brown CW, Groom SB, Miller PI. Analysis of satellite imagery for Emiliania huxleyi blooms in the Bering Sea before 1997. Geophys Res Lett 2003;30:1337. doi: 10.1029/ 2002GL016648 [52] Engelhard GH, Righton DA, Pinnegar JK. Climate change and fishing: a century of shifting distribution in North Sea cod. 2014 Global Change Biology 2014; DOI: 10.1111/ gcb.12513 [53] Winter A, Henderiks J, Beaufort L, Rickaby REM, Brown CW. Poleward expansion of the coccolithophore Emiliania huxleyi. J Plankton Res 2014;36:316–25. [54] Iwataki M, Kawami H, Mizushima K, Mikulski CM, Doucette GJ, Relox Jr JR, Anton A, Fukuyo Y, Matsuoka K. Phylogenetic relationships in the harmful dinoflagellate Cochlodinium polykrikoides (Gymnodiniales, Dinophyceae) inferred from LSU rDNA sequences. Harmful Algae 2008;7:271–7.

Elke S. Reichwaldt, Som Cit Sinang, and Anas Ghadouani

7 Global warming, climate patterns and toxic cyanobacteria 7.1 Introduction “The world is getting warmer.” Most people who read this will think of an increase in the average yearly temperature. However, there is more to it. Heat waves showed unprecedented scales in the last decades [1–7]. More importantly, it is also predicted that their intensity, frequency and duration will increase rapidly under future scenarios of climate change (e.g. [2, 8, 9]). Heat waves can be defined as a number of consecutive days above a certain threshold temperature, usually calculated for a specific location [9]. For example, Australia experienced an extensive heat period in 2013 [4] that led to the introduction of a new colour on the charts of the Bureau of Meteorology (Fig. 7.1). While the effect of heat waves on human health has been studied extensively (e.g. [10]) and governments are starting to develop approaches to coordinate an integrated response to heatwaves to support risk groups (e.g. [11]), the direct and indirect effect of heat waves on lake processes has received very little attention. Screen Temperature Valid 06UTC Mon 14 Jan 2013

ACCESS-Global t+162

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Fig. 7.1: Temperature map of Australia during a heat period in January 2013. Note the purple colour towards the middle, which represents the newly introduced colour code for temperatures between 50 and 52 °C (picture courtesy of the Australian Bureau of Meteorology).

196 | Elke S. Reichwaldt, Som Cit Sinang, and Anas Ghadouani

What do extreme temperature events do to water bodies? Most studies investigated the effect of environmental factors, including temperature, on cyanobacteria on a seasonal or inter-annual scale. However, there is a tremendous lack of information on the immediate and long-term effect of heat waves on cyanobacterial dynamics. Some first information collected during a heat wave in Europe showed that surface temperature, thermal stability and hypolimnetic oxygen depletion increased [12]; the same heat wave was responsible for an increase in cyanobacterial biomass through a combination of higher temperature, reduced wind speed and reduced cloud cover, with the bloom exploding during the heatwave as soon as an artificial mixing process was switched off [13]. Heat waves have also been shown to affect climate patterns, such rainfall patterns [7], which indirectly impact cyanobacterial dynamics. While there is some consensus on the fact that “blooms like it hot” [14], this idea may have led to the generation of a simplistic explanation based solely on an average increase in water temperature. However, there is evidence that extreme events can trigger regime shifts in ecosystems [15–17] and this might also be true for the development of (toxic) algal blooms. This chapter aims to provide a synthesis of the combined effects of direct and indirect consequences of global warming on lake processes and on the occurrence of cyanobacteria; it also presents an outlook of some future scenarios and their implications for society that urgently need further consideration.

7.2 The effect of global warming on inland water bodies 7.2.1 Direct effects of global warming on inland water bodies Global average surface air temperature is predicted to increase by between 1.8 °C and 4.0 °C by the end of this century [18] and this will have severe consequences for inland water bodies [19]. A higher temperature will directly affect lake temperature; this will be especially pronounced in shallow lake systems, because their water temperature usually closely follows air temperature. A higher average temperature or a longer period of warm temperature can affect cyanobacterial growth, bloom development and toxin production directly [13]. Furthermore, any change in temperature-driven lake processes, such as stratification, inflow rates into water bodies, evaporation and nutrient cycling processes will result in changes in cyanobacterial growth (Fig. 7.2), species composition and toxin production. Higher temperatures that prevail for a long period will cause stronger stratifications leading to fast depletion of oxygen in the hypolimnion [12, 20] followed by phosphorous enrichment in the deeper layer [20, 21]. The phosphorous can then be used as an important nutrient source by buoyant cyanobacterial genera that are able to move within the water column (e.g. Microcystis, Anabaena) [13]. Stronger stratification might require events of higher energy to break it down, suggesting that a higher volume of inflow is needed to mix the water column. In addition, higher temperature

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197

Global warming Temperature → water vapour capacity Surface Evaporation runoff

Climate patterns Temperature UV radiation level

Evaporation

Rainfall frequency Rainfall amount

Surface inflow

Stratification Water residence time

Temperature Light

Toxin production

Mixing

Turbidity Conductivity

Cyanobacterial biomass

Phytoplankton Grazers

pH

Subsurface inflow

Nutrients

Fig. 7.2: A conceptual diagram showing the different pathways of how global warming can potentially affect cyanobacterial bloom dynamics. Dark grey arrows depict direct pathways; light grey arrows indicate the indirect pathway through changes in climate patterns; blue arrows indicate the hydrologic pathway.

can lead to a deeper epilimnion and, in temperate lakes, to an earlier onset of the stratification [22], which can lead to extended growth seasons for phytoplankton. Higher temperature in the hypolimnion might also facilitate mineralization of organic matter; this adds to the nutrient pool [23, 24]. A higher temperature will also lead to higher evaporation rates. This can lead to decreased water levels, higher nutrient concentrations and decreased surface inflow into inland water bodies resulting in lower flushing rates and longer residence times, all conditions that favor cyanobacteria [25–27] (Fig. 7.2).

7.2.2 Indirect effects of global warming on inland water bodies Higher atmospheric temperatures will increase the water vapor capacity in the atmosphere [28]. This will have severe impacts on climate patterns, such as the distribution of rainfall on the Earth, the frequency and the intensity of rainfall [28]. These

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predicted changes in precipitation patterns will strongly influence water quality [29]. The symptoms of increased temperature and changes in rainfall patterns are often the same (Tab. 7.1), making it impossible to identify which of the two is responsible for an actual change in cyanobacterial bloom dynamics (Fig. 7.2) (e.g. [30–33]). Tab. 7.1: Correlation between the effects of global warming, specifically an increase in the atmospheric temperature, a change in rainfall intensity and the length of dry periods on water body conditions or mechanisms that affect these conditions, and the direct relationships between these conditions/mechanisms and nutrient enrichment of water bodies, which will then indirectly affect the occurrence of cyanobacterial blooms. Atmospheric temperature

Rainfall intensity

Length of dry period

Condition/mechanism

Nutrient enrichment

+ + + + − + 0

− − 0/− −/+* + − +

+ + + + − + +

+ + 0 0 − +/− +

0

+

−

0

−

+

Water residence time Anoxic conditions Water temperature Conductivity Flushing Water column stability Nutrient concentrations in the inflow after rainfall Turbidity through re-suspension of sediment Turbidity through concentration of biomass

0 0

The table should be read as follows: water residence time is positively correlated to atmospheric temperature, negatively correlated to rainfall intensity and positively correlated to the length of the dry period; water residence time in turn is positively correlated with nutrient enrichment. Explanation for the respective correlations can be found in the text. 0 indicates no correlation. * in areas with highly saline groundwater (e.g. South Australia) [95].

With global warming, it is expected that the total amount of rainfall will increase; however, due to a high inter-decadal variability, the observed trend for the global annual land mean precipitation depends on the period it is calculated from and on the region from which the data was sourced ([28], see also [34]). On a global scale, it is predicted that the frequency of extreme rainfall events will change more dramatically than the mean precipitation rate [28, 35]; heavy rainfall events are predicted to occur more often in the near future while the amount of total precipitation is predicted to change only slowly. The increase in frequency is even likely for regions where a reduction or no change in the total amount of rainfall is predicted [36–40]. This will lead to prolonged dry periods in between events [41], and the probability for droughts is predicted to increase, especially in mid-continental areas during summer [18]. The mean intensity of events has also increased worldwide, and this trend is especially

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pronounced for very heavy and extreme days of rain [39, 42, 43]. Such changes in rainfall patterns will impact the physicochemical conditions within water bodies, which in turn directly affect the ecology of the lake including the occurrence of cyanobacteria (Fig. 7.2). Environmental conditions that are affected by changes in climate patterns include shifts in the concentration of micro- and macronutrients, salinity, turbidity, pH and the stratification of the water body. During rainfall events, water enters the water body directly by wet deposition and indirectly by surface runoff and subsurface flows (Fig. 7.2). The fraction of direct wet deposition of rainwater into water bodies becomes more important the lower the input through groundwater or surface water flow (i.e. small catchment area or distant coastal areas) [44, 45]. While the effect of wet deposition is immediate as rainwater mixes with lake water during the rainfall, the impact of water entering the water body indirectly via groundwater or surface water might take hours to years (e.g. [46]), depending on the size of the catchment area and the geology [47]. The effect of rainfall events on inland water bodies will also differ, depending on whether the lake receives its water primarily through groundwater exchange or through surface streams (e.g. runoff, rivers) [48], because of differences in physical and biological processes in groundwater and surface water that modify, for instance, nutrient species during transport after rainfall events [49]. The effect of rainfall events on water bodies is highly complex and depends mainly on the interplay between the quantity and quality of the inflowing water, the volume ratios of inflowing to receiving water, the seasonal timing of the event and in-lake conditions [34, 50]. The main processes that affect the quantity and quality of the inflowing water in relation to rainfall patterns are the following: the quality of inflowing water depends on the geology and land-use in the catchment area, and the length of the preceding dry period, but also on the intensity of the rainfall event and the chemistry of the rainwater. The quantity of the inflowing water depends on the amount of rainfall, the length of the preceding dry period, the size of the catchment area and the land-use in the area, and is an important driver of water residence time. Although rainfall volume and inflow volume are correlated, this correlation might only be visible if enough rain falls to saturate the catchment [51].

7.2.2.1 Effect of rainfall events on lake temperature, stratification and turbidity Many rainfall events, especially the more extreme ones, occur concurrently with lower air temperatures and strong winds. A significant decrease in air temperature for a prolonged period will especially influence shallow lake systems, because their temperature usually closely follows air temperature. Strong wind events will increase the depth of the epilimnion by mixing and consequently decrease mean temperature in this layer. A high inflow of water can de-stabilize the stratification and cause mixing [52].

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Turbidity, which describes the amount of suspended solids in the water, can increase significantly and for a long period through sediment transported in surface inflows or re-suspension of sediment by storm events [52, 53]. A decrease in the penetration of radiation into the water column has been found after high rainfall [54–56] and during rainy seasons [57]. Furthermore, prolonged dry periods and high temperatures can gradually increase turbidity due to high biomass accumulation. On the other hand, prolonged no-rain periods can lead to lower turbidity and thus higher irradiance penetration in water reservoirs [58] due to a decrease in suspended particles. Another factor that will change due to variation in rainfall patterns is the dissolved organic carbon (DOC) input into aquatic systems. However, the trend is not clear with studies showing either higher or lower amounts of DOC in the water column after longer dry periods (reviewed in [59]). As DOC and especially colored DOC significantly influence the light regime in the water column [60], a change in the concentration of DOC will strongly influence the competition between primary producers and thus affect the occurrence of algal blooms.

7.2.2.2 Effect of rainfall events on water level and water residence time The water level of water bodies fluctuates naturally due to an imbalance of the water budget, which depends on meteorological and hydrological processes such as evaporation rate, precipitation rate, inflow and outflow [61, 62]. In general, prolonged periods without precipitation lead to a decreased water level [63–65] and a recent study indicated that the water level is correlated with the total amount of rainfall per year and the frequency of extreme rainfall events [66]. While the first correlation has a delayed reaction time of weeks to months due to subsurface inflows, the latter is a more immediate effect possibly through an increased surface runoff during high intensity events [53, 66]. Therefore, the predicted increase in the frequency of extreme rainfall events will lead to more variable water levels in the future. This effect might be especially pronounced in closed lakes that rely on precipitation as their inflow, and in reservoirs as the natural outflow is artificially regulated. Lower lake levels generally favor cyanobacteria due to a concentration of nutrients. Water residence time is strongly negatively correlated to the intensity of rainfall [67] and positively to the length of dry periods [64].

7.2.2.3 Effect of rainfall events on lake nutrients Nutrient concentrations (e.g. phosphorous (P), nitrogen (N), iron (Fe)) in water bodies can increase significantly during and after rainy periods [53, 56, 65, 68] leading to eutrophication of water bodies. In general, the processes involved are an input from the catchment (e.g. [69]), suspension of sediment by water inflow and associated wind [53], and an induction of upwelling events of nutrient-rich hypolimnic water [57, 70, 71].

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However, there are different transport mechanisms for N and P with surface runoff and soil erosion being more important pathways for P [72] and the slower flushing of deeper soil waters more important for the flush rate of nitrate-nitrogen (NO−3 –N) [73]. It was also shown that the transport pathways of total P, dissolved P, particulate P, particulate N, and ammonium-nitrogen (NH+4 –N) are different from those of total N, dissolved N, and NO−3 –N, with the first being mobilized quickly once a rainfall events occurs and the release of the latter being slower [74]. As some cyanobacteria seem to be able to take up NH+4 more efficiently than NO3 [75], this might promote an immediate response of cyanobacterial growth after rainfall events. Direct wet deposition of rainwater onto water bodies may have a direct effect on primary production through nutrient input [76, 77]. The addition of N by direct wet deposition to aquatic systems is mainly due to NH+4 [78]. It can be well above the critical loads to systems and can accelerate eutrophication in inland waters, estuaries and coastal waters [79, 80]. Direct deposition of N and P to Lake Taihu (China) by wet deposition alone was calculated as > 4700 and 75 t year−1 , respectively accounting for 16.5 and 7.3 % of the annual N and P input [76, 81]. However, the chemistry of rainwater is highly variable on a regional and seasonal scale [76, 78, 79] with the regional chemical rainfall composition usually reflecting local anthropogenic emission (e.g. agriculture, industry) [45, 76, 79, 80, 82, 83]. The length of the preceding dry period is another important factor shaping the nutrient conditions in inland water bodies. Rainfall events that terminate long dry spells will lead to a comparably higher pulse of nutrients than rainfall events that occur on a regular basis, as nutrients and pollutants build up during the no-rain period (non-urban systems: [84], urban systems: [85]). Additionally, soil moisture, which is affected by the length of the preceding dry period, influences the discharge of N, although this correlation is N species dependent: the highest concentrations of NO−3 –N discharge occurs at relatively low soil moisture while NH+4 –N discharge is lowest at high soil moisture [74]. Also, water soluble organic forms of P are mobilized more easily after longer preceding dry periods or after drying and rewetting due to processes such as microbial or biochemical that facilitate their release [73, 86]. The length of the preceding dry period also affects runoff volume, with larger inflow from dry than from wet catchments [87]. Therefore, disproportionally more nutrients are added to water bodies due to higher runoff volume in dry catchments. Furthermore, phosphorous concentrations in lakes can decrease with longer droughts as the lake retention of nutrients might not be high enough to fully compensate the lower nutrient input due to the loss of inflow streams [60]. Compared to low intensity rainfall events, high intensity rainfall events are able to mobilize and dislodge larger particles and less water soaks into the soil leading to a higher fraction of rainwater input to runoff [88]. A storm with a higher intensity adds more nutrients to a water body than a less intense storm [84, 85, 89] and can lead to massive erosion resulting in very high nutrient input especially into artificial water bodies such as reservoirs [90, 91]. However, the nutrient composition added

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during high intensity events will be biased towards particulate rather than soluble nutrient forms [52, 72, 89], with the former being less available for direct uptake by cyanobacteria. An initial dilution of nutrients in water bodies through the addition of large amounts of nutrient-poor rainwater is in general thought to be negligible as the subsequent input of nutrients by runoff and subsurface flows usually outweighs the dilution effect [74, 92].

7.2.2.4 Effect of rainfall events on lake conductivity and salinity The major processes regulating the conductivity in lakes are soil and rock weathering, atmospheric precipitation and evaporation [48]. This includes also the seasonality of rainfall (summer, winter), the ratio of water entering through surface runoff and groundwater flows, and the conductivity of these flows. The interaction of all these processes, which is a function of catchment characteristics and meteorological parameters, determines the effect of rainfall on the conductivity of a water body [93, 94]. During wet periods, the conductivity of lakes has been shown to depend mainly on precipitation and groundwater inputs, while evaporation controls the conductivity during dry periods [94, 95]. Especially heavy rainfall events can lead to dilution and thus a reduction of the conductivity of the water body [96], because conductivity of rain is lower than that of most lakes [97]. Water loss from water bodies due to evaporation will lead to higher conductivity [64], as the evaporating water contains no or very little ions. However, in areas with high-salinity groundwater (e.g. South Australia), an increase in groundwater inflow to waters could potentially lead to a higher conductivity in the water body [95].

7.2.2.5 Effect of rainfall events on lake pH The effect of rainfall events on the pH of a water body depends on the respective acid neutralizing capacity of the water. The acid neutralizing capacity of poorly buffered systems can be reduced by increased deposition of NO−3 and NH+4 [98] such as during rainfall events, leading to acidic periods during precipitation [99]. An increase in the intensity of rainfall will therefore lead to increased acidic atmospheric deposition [100]. Rain that is in equilibrium with the CO2 in the air has a pH of 5.6 [101], however, emissions from industry can change the pH of rain significantly (“acid rain”). This effect is both local and long-distance and can lead to an acidification of many ecosystems [101–104]. The pH of rainwater can also vary considerably on a seasonal basis (3.3–7.76) [105].

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7.3 The ecology of cyanobacteria and toxin production 7.3.1 Environmental factors affecting cyanobacterial biomass Current studies have proposed a range of physicochemical factors that trigger the occurrence of cyanobacterial blooms in freshwater ecosystems (Fig. 7.3). Excessive nutrient loading and increasing water temperatures are known as the basis for the presence of massive cyanobacterial biomass [14, 106]. As massive cyanobacterial blooms usually occur in eutrophic water bodies, high phosphorus and nitrogen concentrations are assumed to be the primary triggers for their growth [107, 108]. However, the environmental factors causing cyanobacterial blooms in water bodies still remain the subject of a long-standing debate as they seem to be site-specific [109–111] and species-specific (e.g. [112]). Furthermore, eco-physiological differences between cyanobacterial groups in mixed blooms have led to non-homogenous behavior and responses to the natural environment [113]. In terms of phosphorus, Izydorczyk et al. [114] has suggested that high phosphorus concentrations promote high cyanobacterial biomass. On the other hand, de Figueiredo et al. [115] have reported that low phosphorus concentrations favor high

Global warming Climate patterns Cyanobacteria-dominated community Species rich phytoplankton community

ss Ecological factors ma Bio • Light • Temperature • Nutrients • Trace elements • Water residence time • Zooplankton grazing Microcystin production • Competition Genetic level mcy genes

Yes No Toxic strain Non-toxic strain Cellular level Fig. 7.3: Conceptual model of the connection between global warming and climate patterns, and the environmental factors that affect cyanobacterial bloom development and microcystin production.

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cyanobacterial biomass. Similarly, nitrogen was also reported to have either positive [116, 117] or no correlation [118, 119] with cyanobacterial biomass. Evidence has also accumulated that the ratio of total nitrogen to total phosphorus (TN : TP ratio) is important for explaining phytoplankton and cyanobacterial growth [98, 120]; however, this relationship also differs between studies. The hypothesis of low TN : TP ratio favoring the dominance of cyanobacteria [121] has been challenged by studies which reported contrasting findings [122, 123]. The potential correlations between iron concentration in the water column and cyanobacterial biomass have been reported to be either positive [124, 125], negative [126] or have no clear correlation [127]. Higher water temperatures favor the formation of high cyanobacterial biomass through increased growth rates [128], the migration of biomass from sediment into the water column [129], increased stratification and reduced vertical mixing [14], and enhanced hypolimnetic phosphorus accumulation from sediment [23]. In addition, the effects of temperature on cyanobacterial dominance may be direct through its effect on growth rates, or indirect [130, 131], for instance through its effect on water mixing and nutrient transport from sediments [14, 23]. Many studies show a positive relationship between water temperature and cyanobacteria [114, 115, 132, 133]. However, there is also evidence that this is highly species specific and that the length of the warm period is more important than the absolute temperature [134, 135]. Analogously, the correlations between temperature, light and cyanobacterial biomass are also contrasting [115, 131, 135, 136]. In terms of light, it has been suggested that low light availability in the water column is responsible for initiating the presence of high cyanobacterial biomass [131] and especially buoyant cyanobacteria were shown to be favored under low light conditions [58, 108, 137]. However, it has also been suggested that low light conditions were not the triggers for the development of high cyanobacterial biomass, but rather caused by it [75]. The presence of these contrasting results clearly highlights that the development of cyanobacterial blooms is the result of the complex interaction between all environmental factors, especially temperature and nutrients [138, 139].

7.3.2 Environmental factors affecting microcystin production Microcystin is one of the most commonly detected cyanotoxin in inland water bodies worldwide, and its production has been studied extensively. The ability of cyanobacteria to produce microcystin is firstly determined at the genetic level with the presence of microcystin synthesis genes known as mcy genes [140–142] (Fig. 7.3). Microcystin is synthesized non-ribosomally by the thio-template functions of large multifunctional enzyme complexes. Gene clusters encoding the biosynthetic enzyme (mcyS) have been sequenced and characterized in Microcystis genera [140]. It is important to note that over 90 variants of microcystins were identified so far (e.g. microcystin-LR, microcystin-RR) [143], each of which exhibits a different toxicity to organisms [144].

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In nature, cyanobacterial populations may consist of single or mixed species, and a single species may be a mixture of toxic and non-toxic strains [108, 145, 146]. Strains are specific genetic subspecies with slightly different traits [108]. As an example, 11 strains of Microcystis aeruginosa were isolated from a single cyanobacterial population [147]. The proportions of cyanobacterial strains capable of producing microcystin within a single population are highly variable, ranging from 12 % [117, 148] to 80 % [149]. Therefore, the concentration of microcystin produced during a bloom will to some extent depend on variations in the proportion of strains containing mcy genes [117] (Fig. 7.3). In addition, microcystin content in the individual cell can vary considerably in any given strain due to different levels of expression of genes that are involved in microcystin biosynthesis [150]. There is accumulating evidence that toxic strains dominate under higher temperature (reviewed in [151, 152]). A range of environmental factors including temperature [153], light intensities [154–156], phosphorus [157], nitrogen [158], TN : TP ratio [159], iron [160] and the presence of other competing phytoplankton [161] have been suggested to correlate with the level of gene expression involved in microcystin biosynthesis. It is important to note that many studies emphasize the importance of synergistic interactions, specifically of eutrophication and temperature in shaping the toxin production (reviewed in [151]). Further, many studies indicate that toxin production is positively related to cell growth rate [153, 162–164], implying that a higher toxin production rate per cell can be found if growth conditions are improved. However, the role of these environmental factors on microcystin production still remains largely unclear and is inconsistent between studies, partly also due to site and species-specific responses [109–111, 165, 166]. Changes in temperature have been reported to cause up to a three-fold difference in cellular microcystin content [108]. For example, microcystin production increased from 300 to 900 μg g−1 phytoplankton dry mass as water temperature increased from 25 to 29 °C [167]. On the other hand, there are also reports suggesting that microcystin production was reduced from 2 to 1 mg g−1 dry mass when the water temperatures increased from 25 to 30 °C [168]. In general, high concentrations of nitrogen and phosphorus in the water column have been suggested to have a positive correlation with microcystin production [169, 170]. This may be due to extra energy required for toxin biosynthesis in toxic cyanobacteria [159]. For example, higher nitrogen availability was found to be associated with higher microcystin concentrations in non-nitrogen fixing cyanobacteria such as Microcystis aeruginosa [171]. Nevertheless, studies also reported that microcystin production in Microcystis aeruginosa is independent of nitrogen availability [116, 160]. Similar to the role of nitrogen, the effect of phosphorus on microcystin production appears to differ between studies. Wang et al. [172] and Rinta-Kanto et al. [157] have shown that the cellular microcystin content in Microcystis aeruginosa increased with higher total phosphorus concentrations in the water column. In contrast, other studies suggest a negative correlation or no correlation between phosphorus and microcystin production [117, 173]. These conflicting findings are possibly due to non-linear effects

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of nutrients on microcystin production [174]. Graham et al. [174] have demonstrated that microcystin concentrations exceeding 2 μg l−1 occurred when the total amount of nitrogen in the water column varied between 1 to 4 mg l−1 . In contrast, microcystin concentration decreased to below 1.5 μg l−1 when total nitrogen exceeded 8 mg l−1 . Therefore, it can be speculated that nutrient regulation on microcystin production depends on the concentration range of nutrient present in the systems [110]. Complicating this issue further is the fact that the relative availability of N, P, and carbon (C) can lead to the production of different toxins and microcystin variants, because they slightly differ in their stoichiometry [175, 176]. This was hypothesized to lead to water bodies exhibiting different toxicity, depending on the nutrient and light regime [34]. Iron has been identified as an important micronutrient affecting microcystin production [116]. Microcystis aeruginosa was found to produce 20–40 % more microcystin when grown in media in the absence of, or at low iron concentrations (< 2.5 μM ≈ 0.14 mg l−1 ) [177]. It has been suggested that microcystin production is a response to environmental stress associated with iron deficient conditions [178]. On the other hand, contrasting studies find that higher levels of iron (0.9 mg l−1 ) will potentially enhance microcystin production [179]. In addition, results published by Yan et al. [127] have shown that microcystin production was enhanced at higher iron concentrations when cyanobacteria were grown at iron concentrations ranging between 0.5 to 10 mg l−1 .

7.3.3 Ecological factors affecting cyanobacterial blooms: competition Cyanobacteria, like other phytoplankton groups, are a very diverse group with a wide range of physiological requirements. Some general conclusions on the competition between cyanobacteria and phytoplankton groups or other cyanobacterial species have however been made based on the vast number of existing studies. Cyanobacteria are known to have a higher phosphorus uptake rate and lower half saturation constant than other phytoplankton [180]. This means that cyanobacteria can outcompete other phytoplankton under phosphorus-limited conditions. Additionally, cyanobacteria have a higher internal phosphorus storage capacity and can deplete phosphorus to much lower levels than other phytoplankton groups, especially chlorophytes [181]. Therefore, it is possible that low phosphorus concentrations (< 0.1 mg l−1 ) are sufficient to induce excessive cyanobacterial growth [180], and enable cyanobacteria to bloom in oligotrophic waters [108, 182]. Nitrogen fixation is another feature that allows certain cyanobacterial species from the genera Anabaena, Aphanizomenon, Cylindrospermopsis, Nodularia and Nostoc to gain a competitive advantage under nitrogen deficient conditions [108]. Although nitrogen fixation is costly from a physiological perspective, nitrogen fixation is advantageous when the concentrations of dissolved nitrate and ammonium are low [183].

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Cyanobacterial competitive advantages under nitrogen and phosphorus limitation suggest that a low N : P ratio may infer cyanobacterial dominance, particularly of nitrogen-fixing species, over the other phytoplankton communities [64, 121]. Smith [121] has shown that cyanobacterial blooms are likely to occur when the N : P ratio is below 29. However, low N : P ratios might not always determine cyanobacterial dominance as phytoplankton cells have the ability to store a certain amount of phosphorus [184]. Iron is an element with low biological availability and may become a limiting resource for the growth of phytoplankton [185] including cyanobacteria [124]. Cyanobacteria have a competitive advantage to dominate under low iron availability due to their ability to alter their cellular iron requirements and increase their ability to utilize iron at low concentration through the presence of siderophores [184, 186, 187]. For example, many cyanobacterial genera including Anabaena, Microcystis and Planktothrix can produce siderophores to facilitate the uptake of ferric ions into cells under ironlimited conditions [186, 188]. Cyanobacteria are able to absorb light efficiently due to the presence of two reaction centers, PS I and PS II, in their photosynthetic apparatus. This feature allows cyanobacteria to effectively use the light spectrum between the absorption peaks of chlorophyll-a and carotenoids [108, 189]. Additionally, cyanobacteria are also known to have a lower specific maintenance rate than chlorophytes. Hence, cyanobacteria may require smaller amounts of energy to sustain their cellular function and growth [190, 191]. As a result, cyanobacteria are able to maintain a higher growth rate than phytoplankton under low light intensity. Having a competitive advantage in relation to phytoplankton under low light energy supply allows cyanobacteria to become dominant in turbid water [58, 108, 131, 137]. Many cyanobacteria including Microcystis spp. and Aphanizomenon spp. contain gas vacuoles, which give their cells a lower density than water [108]. With the presence of gas vacuoles, cyanobacterial cells are able to control their buoyancy and can migrate vertically along physical, chemical and light gradients to achieve optimum growth conditions [113, 192]. Through this mechanism, it has been suggested that buoyant genera can utilize high phosphorous concentrations in anoxic hypolimnia of water bodies. Also, in stratified water columns, buoyant cyanobacteria have an advantage over heavier phytoplankton, such as diatoms, as the latter will sink to the darker hypolimnion in the absence of mixing events [128]. Shifts between the dominance of different cyanobacterial species are similarly complex and depend on species-specific optimum conditions [193]. A shift between Anabaena and Microcystis was found to depend on the onset of warming, the accumulated heat rate, the dynamics of the stratification and the supply of TN and TP [112]. At this stage, there is no clear evidence to support the theory that microcystin is produced by cyanobacteria in order to provide a competitive advantage in the phytoplankton community. However, microcystin production in cyanobacterial cells is potentially influenced by the interspecific interaction between cyanobacteria with other

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phytoplankton groups, or within the cyanobacterial community. It has been proposed, through the theory of allelopathic interaction, that cyanobacteria may increase microcystin production to outcompete their competitors in ecosystems [161, 183, 194] and that the presence of microcystin in the water column can affect the growth of coexisting organisms such as microalgae [195, 196]. Under growth limiting conditions, the benefit of producing microcystin might therefore outweigh the production cost [149].

7.4 Direct and indirect effects of global warming on cyanobacterial growth As described previously, global warming has profound direct and indirect effects on the physicochemical condition and water quality of inland water bodies [29, 63] (Fig. 7.2). Information on the effects of these changes on cyanobacteria is plentiful, both from field and laboratory experiments. The effect was shown to be highly species and strain-specific and has been discussed and reviewed for many decades [13, 33, 34, 108, 197]. The consensus is that the future is “blue-green” [32], because “blooms like it hot” [13, 14]. In general, cyanobacterial growth is enhanced and cyanobacteria are superior competitors against other phytoplankton at high temperature, under nutrient-rich conditions, at higher salinity and under stratified conditions (Tab. 7.2). However, in particular the effect of nutrients on cyanobacterial growth is partially contradictory, emphasizing that the response will strongly depend on the interaction of different nutrients, and that the competitive advantage of cyanobacteria over other phytoplankton will be highly system-specific [58, 184]. For instance, some cyanobacteria seem to be able to take up NH+4 more efficiently than NO−3 [75], making blooms more likely after rainfall events in catchments with animal husbandry rather than in agriculturally affected catchments [58]. Many studies also emphasize the importance of low N : P or NO−3 –N : TP ratios for the occurrence of cyanobacterial blooms [90, 121, 198, 199]. However, this might be a prerequisite rather than a cause for cyanobacterial blooms as many water bodies with low ratios do not exhibit blooms [64, 121, 200]. This suggests that the N : P ratio is only one factor amongst many that is important for the occurrence of algal blooms. In addition, buoyant cyanobacteria are highly favored during periods of high turbidity and high pH due to their ability to stay close to the surface where they are able to harvest light and CO2 (which is usually low in water with high photosynthesis rates and thus pH > 9) from the atmosphere ([33], cf. [137]). Buoyant genera might also be favored during periods of epilimnetic nutrient depletion due to their ability to utilize nutrient sources from the hypolimnion. In the following we discuss the possible scenarios for cyanobacterial biomass and toxin dynamics in the face of global warming, which is also summarized in Tab. 7.3.

– Rainfall caused mixing and re-suspension of Microcystis cells from the sediment, increased turbidity, high turbulence, nutrient inflow; – Drought caused prolonged periods of high surface water temperature, stable stratification, a decreased depth of mixed layer, nutrient release from sediment in anoxic hypolimnion, long water residence time, reduced or increased turbidity; – High water temperature; accumulated heat rate; stratification; TP supply; TN supply; – Rainfall caused reduced salinity and increased nutrient concentrations; – Drought caused stable water column, high conductivity, low light availability, long water residence time, low wind conditions; an efficient alkalinity system allowed cyanobacteria the use inorganic carbon; – Rainfall caused increased turbidity, high turbulence, de-stratification, lower irradiance, increase in total inorganic N (especially NO3 –N), flushing;

Bloom

Mixed bloom

Decrease in biomass

Cyanobacteria

Potential drivers

Effect

Indicator

Reference [25–27, 53, 56, 64, 65, 91, 106, 212, 263, 264]

[63, 96, 112, 200, 211]

[53, 54, 56, 208, 209]

System type Lakes, reservoir, shallow lakes, rivers, estuaries

Lake, saline lake, reservoir

Lake, shallow lake , reservoir

Tab. 7.2: Summary of studies that investigated the response of cyanobacteria, phytoplankton and ecosystems to increased temperature and rainfall events in inland water bodies.

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Higher diversity

Weaker trophic interactions

Ecosystem

– Flooding destabilizes food webs

– Rainfall caused de-stratification, disturbance;

– Rainfall caused increased turbidity, high turbulence, dilution, mixing, flushing;

– Intrinsic factors more important than environmental factors;

No change in dominance

Decrease

– Appearance of a dinoflagellate bloom due to wet winter;

Interruption of bloom occurrences

– Mixing, higher nutrient concentrations;

– Extreme rainfall caused flushing during intense monsoon years;

Lower frequency and magnitude of blooms

Increase

Potential drivers

Effect

Algal community

Total algal biomass

Indicator

Tab. 7.2 (continued)

Wetland

Shallow lake, reservoir

Shallow lake, pond, reservoir

Lake, oligotrophic lake

Reservoir

Eutrophic lake

Reservoir

System type

[266]

[57, 64, 208]

[51, 53, 57, 63, 265]

[50, 65]

[71]

[50]

[200]

Reference

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Higher toxin production at higher growth ratesa, b ; higher growth rates of toxic compared to non-toxic strainsc CB outcompete Phyto at higher temperaturesa,d,e

Toxin production ↑

CB > Phyto

Lower water level

Higher temperature increases CB growth ratea concentration of nutrients and biomass leads to eutrophicationf and light limitation → CB can adapt to different light levelsg e.g. buoyant CB avoid light limitation by forming surface scumh ↑: higher toxin production at higher growth ratesa, b ↓: non-toxic strains are better competitors for light under light limitationi Higher salinityj , higher temperaturea , higher turbidity due to biomassh ; eutrophicationf

CB biomass ↑

Toxin production ↑ or ↓

CB > Phyto

Low diversity, species-poor system; development of CB bloom with high or low toxicity

Low diversity, species-poor system; development of toxic CB bloom;

Higher growth rates at higher temperaturea

CB biomass ↑

Higher water temperature

Possible consequences for ecosystem

Predicted effects on CB biomass, toxin production and competition between CB and Phyto

Conditions in waterbody

Possible mechanisms

Tab. 7.3: Summary of the predicted effects of likely changes in water body conditions due to global warming on cyanobacterial blooms; ↑ = increase; ↓ = decrease; CB = cyanobacteria; Phyto = non-cyanobacterial phytoplankton; CB > Phyto or Phyto > CB means that conditions favor cyanobacteria or phytoplankton, respectively (adapted from [34]).

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Higher toxin production under growth-favorable conditionsb Stable conditions favor CB that have lower growth rates than Phytoa, d, g, l, m

Toxin production ↑

CB > Phyto

Lower toxin production under unfavorable growth conditionsb Comparatively higher growth rate of other phytoplanktona

Toxin production ↓

Phyto > CB

Eutrophicationf by nutrient input from nutrient-rich deep water layer can rapidly be utilized by CB, especially if epilimnion was nutrient limited beforeg ; re-suspension of CB dormant formsn Light limitation under high C, N supplyo Eutrophicationf

CB biomass ↑

Toxin production ↑

CB > Phyto

Long-term (after mixing conditions)

Unfavorable for CB due to their relatively slow growth ratea

CB biomass ↓

Short-term (during mixing conditions)

Nutrient depletion of the euphotic zone, CB have higher affinity to P,N, Fe than Phytoa ; some buoyant CB species migrate to nutrient rich hypolimnionk

CB biomass ↑

Longer water residence time = stable water column (stratification)

Destabilized water column by water inflow and wind

Possible mechanisms

Predicted effects on CB biomass, toxin production and competition between CB and Phyto

Conditions in waterbody

Tab. 7.3 (continued)

Low diversity, species-poor system; development of toxic CB blooms;

High diversity system; toxin concentration in the system low;

Low diversity, species-poor system; development of toxic CB bloom;

Possible consequences for ecosystem

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Higher toxin production at higher growth ratesa, b ; higher growth rates of toxic compared to non-toxic strainsc Eutrophicationf

Toxin production ↑

CB > Phyto

Increased salinity

Anoxic conditions

Depends on species-specific optimum growth curves Depends on species-specific optimum growth curves CB are more salt-tolerant than many Phytot

Toxin production ↑ or ↓

CB > Phyto

↑: due to favorable growth conditionsb and increased c(Fesol )r , ↓: due to increase of c(Fesol )s

Toxin production ↑ or ↓

CB biomass ↑ or ↓

Nutrients are mobilized from the sedimentq → eutrophicationf

CB biomass ↑

CB possibly dominant but it depends on the salinity level if blooms develop

Low diversity, species-poor system; development of CB bloom with high or low toxicity;

Low diversity, species-poor system; development of toxic CB blooms; higher biomass carrying capacity of the waterbody;

Eutrophicationf by nutrient input through surface and subsurface flows

CB biomass ↑

Nutrient input

Low CB and Phyto biomass levels; toxin concentration in the system low;

Dilutionp

CB and Phyto biomass ↓

Flushing

Possible consequences for ecosystem

Possible mechanisms

Predicted effects on CB biomass, toxin production and competition between CB and Phyto

Conditions in waterbody

Tab. 7.3 (continued)

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Relatively higher growth rate of CB under low light intensitiesa Depends on species-specific optimum growth curves Depends on species-specific optimum growth curves (Buoyant) CB are better competitors at higher pHk

CB > Phyto

CB biomass ↑ or ↓

Toxin production ↑ or ↓

CB > Phyto

d [106], n [70],

CB > Phyto

e [32], f [267], o [176], p [56],

Toxin production ↑ or ↓

Citations are examples only: a [108], b [164], c [153], k [33], l [210], m [26],

High turbidity due to high biomass

g [58], q [90],

h [137], r [268],

i [230], s [269],

j [222], t [223].

Buoyant CB avoid light limitation by forming surface scumh ; CB can adapt to different light levelsg ↑: light limitation under high C, N supplyo , ↓: non-toxic strains are better competitors for light under light limitationi Relatively higher growth rate of CB under low light intensitiesa

↑: light limitation under high C, N supplyo , ↓: non-toxic strains are better competitors for light under light limitationi

toxin production ↑ or ↓

CB biomass ↑

Buoyant CB avoid light limitation by forming surface scumh ; CB can adapt to different light levelsg

CB biomass ↑

Increased turbidity due to sediment re-suspension

Higher pH due to high photosynthesis levels

Possible mechanisms

Predicted effects on CB biomass, toxin production and competition between CB and Phyto

Conditions in waterbody

Tab. 7.3 (continued)

Low diversity, species-poor system; development of CB bloom with high or low toxicity;

CB possibly dominant but it depends on the pH if blooms develop;

Low diversity, species-poor system; development of CB bloom with high or low toxicity;

Possible consequences for ecosystem

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7.4.1 Temperature, stratification, and mixing It is predicted that an increase in average temperature is likely to lead to a higher frequency of toxic cyanobacterial blooms in the future [13, 100]. Increasing water temperatures are known as the basis for the development of massive cyanobacterial biomass [14]. Temperatures exceeding 25 °C are likely to provide optimum conditions for cyanobacteria to grow rapidly and dominate in water bodies, through regulation of cyanobacterial photosynthetic capacity, respiration and growth rate [130]. Earlier studies have shown that the optimum temperature for cyanobacteria is higher than for chlorophytes and diatoms, indicating that global warming will favor cyanobacterial dominance in the phytoplankton community (see reviews in [108, 193]) leading to lower algal diversity [64]. However, more recent work did not find differences in the optimum temperature for cyanobacterial and chlorophyte growth; it was concluded that the ability to dominate was rather due to the cyanobacteria being able to avoid sedimentation during long stratification periods [201]. In any case, the direct effects of temperature on cyanobacterial dominance will be intertwined with the effects of other environmental factors (Tilman et al. 1986, as cited in [130]), making any prediction of bloom development and therefore management of the water body extremely complex. Global warming has been shown to be responsible for a wider geographical distribution of cyanobacteria worldwide. The best known example is the invasion of Cylindrospermopsis raciborskii, a tropical and subtropical species into temperate water bodies (e.g. [202, 203]. It has been suggested that this species is able to extend its current distribution due to higher average water and sediment temperatures and due to an ongoing eutrophication of many water bodies [204]. More recently, it was also suggested that the distribution of Microcystis could expand under future climate scenarios [166]. The higher stability of the stratification during periods of higher temperature and no rain might favor buoyant cyanobacterial genera such as Anabaena, Aphanizomenon and Microcystis as they are promoted during non-mixing phases [112, 205, 206], although Oscillatoria growth was shown to be stimulated by mixing events [207]. In regions with distinctive dry seasons, blooms prevail longer if the start of the rain is delayed or the rainfall events are too weak to break up the stratification [58]. Higher temperature and long dry periods can also increase the water retention time, and cyanobacteria were shown to dominate during periods of low flushing [26, 27]. The inflow of high volumes of water during rain events can lead to a direct reduction of algal biomass due to high flushing rates [51, 54, 56, 57, 63, 208] (Tab. 7.2) and can significantly change the community composition possibly due to persistent high turbidity and turbulent conditions which favor diatoms and small-celled cyanobacteria over large-celled or filamentous cyanobacteria [53]. Additionally, heavy rainfall events can lead to a mixing of the water column [56, 63, 209], which might favor noncyanobacterial species [63]. Thus, in the short term, intense rainfall can lead to a lower total chlorophyll biomass with a higher diversity due to the absence of cyanobacterial

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dominance [64]. A movement of the thermocline by water inflow or winds associated with storms can also allow nutrients that are released from sediments to enter the photic zone, leading to an additional nutrient enrichment and can re-suspend cells (e.g. Microcystis spp.) or dormant forms of cyanobacteria (e.g. akinetes from Cylindrospermopsis spp.) into the water column which encourages bloom development [65, 70].

7.4.2 Nutrients In general, increased atmospheric temperature can lead to eutrophication supported by lower water levels, longer water residence time, an increased release of nutrients (e.g. soluble P, NH+4 , Fe) from the sediment if anoxic hypolimnia develop [90] and increased mineralization rates at higher temperature [23, 24]. The nutrients released from the sediment can be utilized by buoyant cyanobacteria that are able to control their vertical movement in the water column [135] and additionally contribute to the eutrophication process during mixing events [108]. All these conditions will strongly favor freshwater cyanobacteria [56, 63, 210] and it was demonstrated that cyanobacterial blooms dominate during drought and falling water levels [51] and that surface scums can occur during dry periods with low wind velocities [211]. Less intense rainfall events can immediately increase cyanobacterial biomass through nutrient enrichment, if the event does not lead to de-stratification as is often the case for isolated rainfall events or in shallow, non-stratified lakes [58, 212]. The importance of nutrient addition to inland water bodies by direct wet deposition can be significant [76, 79, 81]. Especially in closed lakes, nutrient addition through precipitation is the main source for phytoplankton [213]. However, there is still little understanding of the magnitude of this impact, especially in the context of cyanobacterial bloom development. One study indicated a clear connection between the addition of N from wet deposition and the occurrence Microcystis [214]. Also, the TN : TP ratio of wet and dry deposition has been identified as an important factor to drive the TN : TP ratio and phytoplankton biomass in alpine lakes [215], and it was shown that dominance of non-nitrogen fixing cyanobacterial species (e.g. M. aeruginosa] depend more on the TN : TP ratio than on TN alone [108, 216]. Therefore, an addition of N through precipitation will also affect systems that are not N-limited by definition, such as many inland water bodies, and it has been emphasized that N can play a major role in eutrophic [217–219] and tropical freshwater systems [220]. This is supported by a study that suggested that atmospheric deposition of N and P might have promoted cyanobacterial blooms during periods that are in other respect optimal for their growth [221]. Prolonged periods of wind which are related to more intense storms or longer periods of rain can enhance the effect of nutrient input during rainfall by mixing events for up to three months (Bormans et al. 2001 in [58]), and the large nutrient input during intense events will lead to nutrient-rich conditions which generally favor cyanobacteria.

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7.4.3 Salinity In general, higher atmospheric temperature can lead to increased salinity due to high evaporation rates. An increased salinity might selectively favor cyanobacteria, as many species are more salt tolerant than other phytoplankton species. For instance, Microcystis aeruginosas’ long-term growth was not affected by salinity up to 9.8 g l−1 [222] with short term tolerance to salinities as high as 17.5 g l−1 [223]. In contrast, rainfall events can decrease salinity [96]. Therefore, the first period after a rainfall event in which salinity is decreased only slightly while there is already a nutrient pulse might represent an important timeslot that gives cyanobacteria a competitive advantage.

7.4.4 Turbidity and pH Cyanobacteria and specifically buoyant genera are highly adaptive to different light levels, and higher turbidity generally favors cyanobacteria [137]. Sudden low light availability after a rainfall event, in combination with turbulence, was shown to lead to a decrease of large-celled cyanobacterial species [53] (Tab. 7.2). Increased light availability was shown to correlate well with the biomass of a Microcystis aeruginosa bloom in the presence of a sufficient nitrogen source [25], while a decrease in nitrogen was shown to favor N-fixing cyanobacteria [224]. If the higher turbidity is the result of algal growth, this can lead to a higher pH due to increased photosynthesis; in turn, this would limit dissolved inorganic carbon (DIC) in the water column. Both conditions would favor buoyant cyanobacterial species (e.g. Microcystis, Anabaena, Planktothrix) over other phytoplankton as their ability to dwell directly on the surface (scum) enables them to harvest light and CO2 from the atmosphere thus avoiding DIC limitation [137].

7.5 Direct and indirect effects of global warming on microcystin concentration There are no studies that directly investigate changes in cyanotoxin concentration in water bodies following rainfall events or prolonged periods of heat. However, plenty of information on the drivers of toxin production is available from laboratory studies. Furthermore, numerous field studies tried to correlate in-situ toxin concentrations to environmental conditions. All these results can be used to draw conclusions on how toxin concentrations might change with global warming. Cyanotoxin concentrations in a water body depend on the toxin content within each cyanobacterial cell, the biomass of toxic cyanobacteria and on toxin-degrading processes. Therefore, it is critical to understand the possible effects of global warming

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on all these processes. Cyanobacterial toxin concentration in the water is also a function of the succession of species and strains. Conditions leading to species-specific proliferation of cyanobacterial cells is fairly well understood; for instance Oscillatoria can outcompete other phytoplankton and cyanobacteria under mixed, low light conditions, while Anabaena, Aphanizomenon and Microcystis will dominate under stable conditions with higher underwater irradiance [206]. In contrast, the factors that lead to the dominance of toxic or non-toxic strains are still relatively unknown [225]. It can be hypothesized that toxin concentration in the water will be higher under future rainfall scenarios and global warming due to an increase in cyanobacterial biomass alone [226]. However, higher biomass does not necessarily mean that growth conditions (and thus the toxin production rate) are optimal for cyanobacteria. Therefore, whether global warming will lead to higher or lower average toxin concentration in a water body will ultimately depend on the effect of temperature and rainfall patterns on (variant-)specific toxin production rates [34]. As discussed previously, the toxin production per cell is a function of the environmentally influenced gene expression [108]. A range of environmental factors including temperature [153] have been suggested to correlate with the level of gene expression involved in microcystin biosynthesis. As with the effect of temperature on cyanobacterial growth, the direct effect of temperature on toxin production seems to be species and strain specific, and is intertwined with the effect of other factors (reviews: [108, 162]). As changes in temperature have been reported result in an increase [167] or decrease [168] of the microcystin production, no simple prediction can be made on the effect of increased water temperature on the toxin concentration in the water body. However, there is accumulating evidence that toxic strains dominate under higher temperature (reviewed in [151, 152]) emphasizing that the future might hold more toxic blooms. Decreased turbidity following a heavy rainfall event or due to high biomass during periods of high temperature and/or no rain can lead to higher light penetration; this might increase microcystin production [227]. The second important factor affecting the toxin concentration in the water is the biomass of toxic cyanobacteria [150]. Very often, total cyanobacterial biomass is used as an indicator of the toxin concentration in the water column [228]; this works well if a bloom consists of single species and strains. However, cyanobacterial blooms often contain a mix of toxic and non-toxic species. Even those species that are potentially toxic are comprised of toxic and non-toxic strains (genetic subgroups of species) [108, 229] that might have different optimum growth conditions. For instance, nontoxic strains of M. aeruginosa were shown to be better competitors for light than toxic strains [230], while higher water temperature and higher concentrations of inorganic N and P are likely to promote toxic strains ([219], reviewed in [151]). It has also been demonstrated that the succession of different strains was responsible for high variations in toxin concentration in the water over the lifetime of a bloom [231]. Therefore, any predictions of the effects of global warming on the biomass dynamics of toxic cyanobacteria have to be based on our understanding of toxic and non-toxic species

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and strain succession, especially as there is evidence that the toxin production rate per cell is often positively correlated with cyanobacterial growth [153, 162–164]. However, little is known in this respect and results are highly ambiguous, indicating that favorable growth conditions for cyanobacteria can lead to the dominance of non-toxic or toxic strains [149, 150, 153, 163]. The third mechanism that influences the total toxin concentration in the water includes processes by which the concentrations of dissolved toxins are decreased [232]. These processes include dilution, biodegradation by bacteria [233, 234], adsorption onto particulate organic matter and sediments [234–236], thermal decomposition and photolysis [237]. No direct study of the effects of rainfall events on these processes has been performed, but it can be inferred from the literature that increased water temperature through higher atmospheric temperature or prolonged dry periods with high evaporation enhances thermal biodegradation and increases bacterial biodegradation [238]. Heavy rainfall events that input large amounts of sediment and particulate organic matter and/or stirs up sediment could increase the amount of adsorbed toxins to sediment or alternatively release loosely bound toxins into the water column [239]. The input of large water volumes could dilute the dissolved toxin fraction, and changes in pH can lead to changes in the adsorption of toxins onto sediments, with a decreased sorption at higher pH [240]. Most of these processes would decrease the concentration of dissolved toxins in the water column. However, as the dissolved toxin fraction is usually small compared to the intracellular fraction, the impact of possible changes with global warming might be less relevant than mechanisms that influence the intracellular fraction.

7.6 Why should we care? How are we responding to the cyanobacterial problem in its new context of climate change? Cyanobacterial blooms have been around for a long time [241] and, since then, there has been an enormous scientific interest in understanding the dynamics of cyanobacteria and their harmful compounds. Cyanobacteria have been considered as a “nuisance” [210] that is harmful to aquatic organisms and humans [114] and, as such represent a “global public health threat” [242]. This potential hazard is reflected in numerous guidelines adopting critical values for accepted toxin concentrations in drinking and recreational waters [225, 228, 243]. These guidelines are an important response to minimize cyanobacterial-related health issues especially in the light of global change that is very likely to increase cyanobacterial occurrence. More recently, awareness has been raised that apart from a health related impact (e.g. [244]), cyanobacterial blooms have implications for many other sectors, such as the water resource sector, infrastructure, policies and the social sector. It is important that this awareness is reflected in public documents to ensure an adequate reply to this increasing problem.

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An extensive series on the “Perspectives on water and climate change adaptation” by the World Water Council, the Cooperative Program on Water and Climate and the International Water Association in cooperation with other contributors was a valuable contribution to identify areas of impact on water by climate change with the aim to “define and distil the critical role of water in climate change adaptation and to lay out strategic and operational priorities for adaptation of water management and services” [245]. In the included papers, water quality was identified as a serious problem for food production [246], the water industry [247] and WASH-service delivery [248]. Interestingly, cyanobacteria were only mentioned once, indicating that the link between industry and science is not complete. This is of particular concern as “information and data collection and sharing” has been identified as an extremely important strategy to adapt to climate change in all water-use sectors [245]. Furthermore, the IPCC [100] acknowledges that adaptation procedures including risk management practices are being developed by the water sector in some countries, but this is mainly directed to hydrological changes [249]. The implications of the deterioration of water quality, including cyanobacterial blooms by climate change for the water sector, are described as being substantial but seem to be less developed. The implications of cyanobacteria on our life are manifold (Fig. 7.4) and the intensity of the impacts can be expected to multiply in the future. While the direct implications on the water resource sector, including wastewater and drinking water treatment, are obvious, the impact of cyanobacteria on the infrastructure sector, the food sector and the society are less explicit. Cyanobacterial blooms are now regularly found in drinking water reservoirs (e.g. [250, 251], which increases the likelihood of the presence of toxins and taste and odor compounds in this water resource [27]). Their presence incurs high costs for infrastructure maintenance, treatment of the water and for provision of alternative drinking water to the community during blooms. New, costeffective treatment methods for the removal of toxins (e.g. [252]) and taste and odor compounds are needed to counteract increasing costs in the future. Further, an increase in the occurrence of cyanobacterial blooms during wastewater treatment processes will hinder to achieve high water quality effluent that can be reused for multiple purposes, including irrigation. Similar to the problems with drinking water, costefficient treatment methods have to be developed (e.g. [253–255]) to allow the safe future use of this important water resource. The occurrence of cyanobacteria will certainly also impact our food security due to less clean water being available for irrigation and watering of animals. It will make agricultural products more expensive due to higher costs for treating the water before use or transporting good quality water. Aquaculture and fisheries will also be impacted by the future increase in cyanobacteria, both in inland water bodies and the ocean [256–260]. The presence of blooms will lead to shorter fishing seasons. This, in combination with the necessity to conduct more rigorous monitoring of toxin concentrations in seafood, can lead to enormous economic losses [260].

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One often-forgotten ecosystem service that water delivers is its role as a recreation source [261]. The recreational value of lakes, rivers, estuaries and the coast will very likely suffer in the future. A recent study showed that in some countries the recreational use of lakes is at risk due to cyanobacteria and high nutrient levels in as many as 50 % of the studied lakes [262]. This indicates not only a huge impact on recreation, but also large economic losses due to the cleaning of infrastructure and water equipment, such as boats and boating harbors’ infrastructure. This is not only expensive but potentially also harmful. Additionally, tourism and recreational fishing have been shown to be negatively affected by blooms (e.g. [244]) and this is likely to increase even further in the future.

Society Policies

Policies Public health

Treatment costs

Treatment costs

Drinking water

Health Waste water

Infrastructure

Infrastructure Toxic cyanobacterial bloom

Reuse

Policies Public health Treatment Environment costs Infrastructure

Public health

Recreation

Food Infrastructure

Policies Treatment costs

Fig. 7.4: Important implications of cyanobacterial blooms for the society.

In summary, toxic cyanobacterial blooms are directly and indirectly impacting society, making this a global problem that is likely to worsen in the future. So, where do we go from here? There seems to be sufficient scientific evidence that the occurrence of toxic cyanobacterial blooms will very likely increase in many water bodies and regions world-wide due to a combination of climate change, eutrophication and manmade changes of the hydrology of water bodies. Although the exact bloom dynamics will certainly be site-specific, the general patterns seems to be clear. Therefore, it is

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now time to take the next logical step and focus on defining the implication of toxic cyanobacterial blooms for the public, for the governments, for businesses and most importantly for the water sector. This requires a close collaboration between social and natural sciences and will need substantial support by the private sector, politics and stakeholders. Just like the dynamics of cyanobacterial blooms is a complex issue, solving it will have to be a collective effort. And only when these implications have been clearly identified and perceived by the public can we start developing clear strategies to adapt to a blue-green future.

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Aristidis Vlamis and Panagiota Katikou

8 Human impact in Mediterranean coastal ecosystems and climate change: emerging toxins 8.1 Introduction An increase in the occurrence of harmful or toxic algal incidents, both in frequency and in geographic distribution, has been observed in many parts of the world over the past three decades [1–3]. The predicted changes in the oceans have both a direct and indirect impact on interactions between humans and the oceans, and over the past decade, numerous studies have suggested possible relationships between climate and the magnitude, frequency, and duration of HABs [4]. Dense phytoplankton blooms develop in response to favorable conditions for cell growth and accumulation [5]. These blooms of autotrophic algae and some heterotrophic protists are increasingly frequent in coastal waters around the world. Undoubtedly, HABs are occurring in more locations than ever before; new sightings are reported regularly. Certain researchers argue that this trend is due to increasing eutrophication throughout the world [6] but, generally, phytoplankton blooms have regional, seasonal and species-specific aspects that should be considered [7–9]. In contrast to large-scale blooms that are dominated by mesoscale circulation, Mediterranean HABs are a more localized phenomenon commonly present in areas of constrained dynamism, such as bays, lagoons, ports, beaches and estuaries. In such areas, enhanced growth of phytoplankton can cause an observable water discoloration along the shoreline as well as deterioration in water quality. In recent years, certain unprecedented ecological effects on the Mediterranean area, such as fish kills and risks to human health, have also been attributed to toxic algal proliferations. Taking into account that a bloom represents a deviation from the normal biomass cycle, and despite the fact that in some cases algae proliferation may be of a natural origin, coastal blooms are considered an emerging problem possibly related to nutrient enrichment of coastal waters. Furthermore, intensive urbanization and recreational use of coastal watersheds has led to a remarkable increase in nutrient sources along the Mediterranean coasts. This cultural eutrophication creates a contrast between coastal waters and the open ocean where, due to summer stratification and nutrient depletion, oligotrophic conditions prevail in the upper layer. Nutrient-rich coastal environments of the Mediterranean Sea and, in particular, semi-enclosed areas with low turbulence levels constitute a new and unique environment in which a number of phytoplankton species with harmful effects can become dominant [10, 11]. Despite the fact that most of the factors implicated in the Mediterranean nearshore algal outbreaks are known, the mechanisms underpinning their occurrence are not yet well established [12]. Along North African coasts, the spatial distribution

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of chlorophylls and carotenoids is attributed to the human-altered patterns of the physical structure and the nutrient concentrations, as well as to the Modified Atlantic Water (MAW). The physical forcing resulting from advection of the MAW could confront distinct water masses and result in potential mixing of water from coastal and/or open-ocean origin. This water mixing may affect phytoplankton populations, which, in North Africa, present large variations in terms of abundance, composition and size structure due to the dynamic nature of their environment [13]. In the Black Sea, since the late 1970s, anthropogenic nutrient enrichment has been recognized as a key ecological problem for this basin, especially its north-western and western parts, which are mostly subjected to the influence of freshwater nutrient inputs. Nutrient input and dissolved organic matter in the northwest shelf of the Black Sea by the Danube, the Dniepar and the Dniestar rivers has showed an increase of about 10 times in the period 1950–1980. A rise in phytoplankton blooms’ frequency, involved species, duration, timing and area has been documented, provoking substantial perturbations of the entire food web structure and functioning. Changes in zooplankton communities’ structure and deterioration of benthic coenoses, culminating during the 1980s (period of intensive eutrophication in the Black Sea), were to a large extent connected with dramatic alterations in phytoplankton communities and recurrent hypoxic conditions. Microalgal blooms were therefore recognized as one of the major issues for the Black Sea’s ecological health. Similar eutrophication problems have been identified in the Eastern Mediterranean Sea in several Aegean and Ionian coastal areas, affected by urban and industrial wastewaters and/or nutrient inputs from rivers and agricultural activities. In this context, phytoplankton, as a primary producer, became the first target of anthropogenic-induced stress, resulting in dramatic alterations in species composition, abundance and biomass, seasonal dynamics and succession in the two basins [7, 9].

8.2 Mediterranean coastal ecosystems The Mediterranean Sea constitutes the crucial environmental factor of the Mediterranean region. The presence of a large marginal and almost completely closed sea on the western side of a large continental area is geographically unique. Its size is substantial, with an area, excluding the Black Sea, of about 2.5 million km2 , an extent of about 3700 km in longitude and 1600 km in latitude and an average depth of 1500 m. The Strait of Gibraltar, connecting the Mediterranean Sea to the Atlantic, has a width of only 14.5 km and a depth of less than 300 m at the shallowest sill. These morphological characteristics make the Mediterranean Sea a large source of moisture and a heat reservoir with a significant capacity for the surrounding land areas (considering the annual average, it acts as a moderate source of heat). The Strait of Gibraltar has a particular role in the Mediterranean Sea environment. The fluxes through the strait compensate for the mass deficit due to the large evaporation in the basin, supply

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comparatively freshwater masses to one of the saltiest seas on Earth and also provide a relatively small supply of heat because the Mediterranean water (MW) outflow is cooler than the Atlantic water (AW) inflow. The region’s complicated morphology – with the presence of many sharp orographic features, often close to the coastlines, and of distinct basins and gulfs, islands and peninsulas – has a strong influence on atmospheric circulation. Moreover, it is responsible for several cyclogenetic areas, local winds, many mesoscale processes and intense air–sea interactions, such as those responsible for dense-water formation processes driving the Mediterranean thermohaline cells. Furthermore, the shape of the Mediterranean Sea bottom, with deep basins linked through much shallower straits, strongly constrains Mediterranean Sea circulation [14]. Despite the fact that the Mediterranean is a semi-enclosed sea with a relatively small area compared to other large marine ecosystems worldwide, its waters and coasts are characterized by a disproportionately huge diversity from the environmental and socio-economical points of view. Three continents (Europe, Asia, Africa) and 21 countries surround this basin, a number which on its own indicates the great variety in human culture and socio-economic development found in Mediterranean coastal zones that have been inhabited for millennia. Marine environmental and ecological conditions as well as habitats present a great variation: a wide range of depths is covered, and primary productivity displays a clear west–east and north– south decrease, similarly to fishery landings. It is a spatial pattern opposite to that of temperature and salinity [15–18]. The natural balance of Mediterranean coast ecosystems is severely affected by major disturbances, which could cause an extensive loss of biodiversity. Such disturbances include habitat destruction and alien species introductions, overfishing and pollution; these phenomena are tightly related with the increase of anthropogenic pressure due to urbanization on the northwestern shores on the one hand, and outstanding growth population on the southern and eastern shores on the other. The Mediterranean basin is therefore of major interest for studying climate change effects [19]. A large number of coastal areas in the Mediterranean Sea are eutrophic, a situation that is more intensely present in semi-enclosed areas [20]. The use of satellite imagery on chlorophyll distribution has shown that the highest concentrations are located close to river deltas and estuaries or near urban agglomerations, especially in the estuary of the Nile (from Alexandria to Gaza), Gulfs of Antalya and Alexandretta (Turkey), Northern Aegean, Thermaikos Gulf (Greece), the Adriatic Sea, the Gulf of Lions (France), Valencia-Barcelona (Spain) and the Gulf of Gabes (Tunisia). Increased temperatures associated with eutrophication can enhance the occurrence of harmful algal blooms (HABs), negatively impact aquaculture production (especially filter feeders farming) and increase human health risks. In fact, human consumption of seafood contaminated with harmful biotoxins can result in a variety of commonly known intoxications with varying degrees of severity, including amnesic shellfish poisoning

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(ASP), shellfish poisoning from lipophilic toxins (such as diarrhetic shellfish poisoning, DSP) and paralytic shellfish poisoning (PSP), whereas the increased presence of new/emerging marine biotoxins of unknown human toxicity further complicates the problem. Marine biotoxins primarily target the nervous or digestive systems, but can also result in potentially fatal acute respiratory distress and other chronic neurological and immunological illnesses. The danger of inadvertently consuming biotoxins is compounded by the fact that they lack odor and taste, and are unaffected by food preparation procedures [21].

8.2.1 Human impact A worldwide proliferation of harmful (toxic, food-web altering, hypoxia-generating) algal blooms (HABs) has been linked to human nutrient (phosphorus (P) and nitrogen (N)) over-enrichment. Human activities are probably the strongest drivers of change in marine biodiversity at all levels of organization; hence, future trends will depend largely on human-related threats [22]. Most human activities present a local impact, whereas others and, above all, the overall total – acting synergistically – may have global impact through cumulative processes. Building of the Aswan High Dam in 1968 may not only have deleteriously affected the productivity, biochemistry and food web structure in the Nile delta and Eastern Mediterranean, but also the hydrological functioning and structure of the Mediterranean as a whole, which itself will influence the chemical and biological characteristics in a feedback loop [23, 24]. Similarly, one of the most important anthropogenic effects was the opening of the Suez Canal in Egypt in November 1869, which was the start of a massive invasion of hundreds of marine alien species from Indo-Pacific origin [25–28]. With the opening of the Suez Canal in 1869, two markedly different zoo-geographical areas were joined: the subtropical Mediterranean Sea, which connects with the Atlantic; and the tropical Red Sea, the most northern extension of the Indian Ocean. In order to pass between these areas, organisms must be able to bridge the difference in adaptive requirements and also withstand the extreme conditions in the Canal itself [29]. In this context, the term “Lessepsian migration”, named after Ferdinand de Lesseps, the French diplomat in charge of the canal’s construction, has been introduced to characterize a new phenomenon of unidirectional and successful biotic advance from the Red Sea to the Eastern Mediterranean (Fig. 8.1), while the term “Lessepsian migrant” refers to the Red Sea species that have passed through the Suez Canal and settled in the Eastern Mediterranean [30]. The Lessepsian migrant Lagocephalus sceleratus (Gmelin, 1789), also known as silverstripe blaasop, is an Indo-Pacific originated pufferfish of the family Tetraodontidae, which is now very well established in the Eastern Mediterranean. Similar to its congeneric tropical species, L. sceleratus may be a source of food poisoning with a high associated risk of mortality, as it commonly contains tetrodotoxin (TTX), a toxin which can cause death by muscular paralysis, respiratory depression

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Fig. 8.1: Geography of the Mediterranean Sea with the main routes of species range expansion. Bold capital abbreviations correspond to the main Mediterranean subregions (ALB: Alboran Sea; NWM: North Western Mediterranean; TYR: Tyrrhenian Sea; ADR: Adriatic Sea; ION: Ionian Sea; AEG: Aegean Sea; LEV: Levantine Basin) and adjacent seas (ATL: Atlantic Ocean; BLA: Black Sea; RED: Red Sea). Italic abbreviations correspond to some remarkable Mediterranean locations (Gib: Gibraltar Straits; GoL: Gulf of Lions; Sue: Suez Canal). Reported temperatures correspond to winter–summer mean sea-surface temperatures. Arrows represent main routes of species range expansion according to their origin: Mediterranean natives (orange), Atlantic migrants (green) and Lessepsian migrants (red). Reprinted from Trends in Ecology & Evolution, Vol. 25(4), Lejeusne C, Chevaldonné P, PergentMartini C, Boudouresque CF, Pérez T, Climate change effects on a miniature ocean: the highly diverse, highly impacted Mediterranean Sea, Pages 250–260, Copyright (2010), with permission from Elsevier.

and circulatory failure [31–33]. The presence of this alien species and its associated toxin (TTX) in the Mediterranean Sea in the last decade therefore constitutes a new and highly important emerging risk that has to be taken into account by the authorities of the affected countries. Furthermore, around half of all the extra CO2 produced so far by human activities has dissolved in the oceans. The anthropogenic increase in atmospheric carbon dioxide concentration results in a reduction of seawater pH on a global scale (a process termed as “ocean acidification”), changing the chemistry and bio-geochemical cycling of carbon and carbonate. A growing number of studies have demonstrated the adverse effects of acidification on marine organisms. The combination of elevated temperature and acidification has been proved detrimental to the calcification process, hence marine organisms with calcareous skeletons, shells or plates are expected to experience problems. These include major bio-constructing organisms, such as scleractinian corals, bryozoans or red coralline algae, but also mollusks, crustaceans, echino-

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derms, foraminifera and some calcifying phytoplankton (mainly coccolithophorides) as well as zooplancton (pteropods and larvae of many groups) [24]. Mediterranean coastal ecosystems were and are still undergoing heavy human-derived impacts [34]. Large concentrations of contaminants have accumulated in sediments of coastal lagoons due to past industrial activity, thus one of the important issues to be faced is the fate of these contaminants with changing environmental conditions, including those related to climate change [19, 35].

8.2.2 Socio-economical implications of Climate Change Coastal marine systems are among the most ecologically and socio-economically vital at a global level. There is strong scientific consensus for the fact that these marine ecosystems, along with the resources and services they provide, are significantly threatened by anthropogenic climate change [36, 37], with HABs and invasive marine species being within the major issues creating problems, especially in terms of the presence of emerging toxins. The HAB designation is mostly a societal concept rather than a scientific definition – blooms are considered to fit the HAB criterion if they cause injury to human health or socioeconomic interests, or to components of aquatic ecosystems. Some HAB species are toxigenic and produce blooms that cause illness and death of fish, seabirds, mammals and other marine life, often via toxin transfer through the food web. Human consumers of seafood contaminated by these toxins may also be poisoned, suffering acute toxic symptoms and even fatalities in extreme cases. Further toxic threats to human health are posed by toxic aerosols and waterborne compounds that cause respiratory and skin irritation when released from toxic cells [38]. The economic effects of HABs arise from public health costs including morbidities and mortalities, commercial fishery closures and fish kills, declines in coastal and marine recreation and tourism, as well as the costs of monitoring and management. Aggregating economic effects both within and across these categories can be problematic, as the measures of effects are rarely the changes in economic surpluses sought by economists [39]. A rough estimate of the economic effects of HABs in the United States is $ 100 million per year (at the 2012 value of the dollar). Anderson et al. (2000) estimated the proportional breakdown of costs related to HAB impacts to be: 45 % for public health costs; 37 % in term of the costs of closures and losses experienced by commercial fisheries; 13 % to the impact on lost recreation & tourism; and 4 % to monitoring and management costs. As far as the European Union (EU) is concerned, the socio-economic impact of HABs for three evaluated Mediterranean countries – Italy, Greece and France – has been estimated at around 329 million Euro per year [40]; however, approximately two-thirds of this has been associated with the noxious, but non-toxic, effects of macroalgal (and some microalgal, e.g. Phaeocystis) blooms affecting the human uses of the coast [41, 42].

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On the other hand, the significance of invasive species (also known as alien, exotic, nonindigenous, introduced or non-native species) in marine ecosystems worldwide has been highlighted and discussed intensely in recent years from ecological, environmental and economic points of view [43–47]. Some of these species can become economically harmful and sometimes even threaten human health [27, 48], like the Lagocephalus sceleratus pufferfish which has been ranked among the 100 ‘worst’ Invasive Alien Species (IAS) in the Mediterranean Sea with profound social and ecological impacts [49].

8.2.3 Effect to ecosystem from extreme events of climate change Impacts of climate change on the Mediterranean environment particularly relate to water, via a change of its cycle due to a rise in evaporation and a decrease in rainfall. Extreme events, such as heat waves, droughts or floods, are likely to be more frequent and violent. A significant decrease in rainfall, ranging between −4 and −27 % for the countries of Southern Europe and the Mediterranean region is expected (while the countries of Northern Europe will report a rise between 0 and 16 %) [50]. Concomitantly, an increase in drought periods related to high frequency of days during which the temperature would exceed 30 °C is also predicted [51, 52]. This water problem is expected to be of crucial importance with regard to the issue of sustainable development in the region. For instance, floods (resulting from altered rainfall patterns) will affect nutrient loads in the coastal aquaculture areas. High inorganic sediment loads can reduce or arrest the filtration rates of bivalves. Elevated nutrient levels can also stimulate the evolution of HABs. For coastal and offshore aquaculture, more frequent and intense storms result in increased physical damage and stock losses, both of which are costly to operations. Many coastal processes, such as sediment transport, happen mostly during high-energy events (storms). An increase in storm activity may therefore initiate erosion. Any severe flooding event could result in mass mortalities of animals in aquaculture ponds, open-water rafts and lines or cages in coastal and offshore areas. Regarding drought, over much of the Mediterranean basin the general tendency is towards decreasing rainfall. The predicted water stress is thought to result in decreasing water availability in the major Mediterranean freshwater systems, areas where there are important aquaculture activities [21, 24]. The effects of climate change on seafood safety are a relatively new topic. Currently, the issue of seafood vulnerability to climate change is scarcely considered both at national and at international levels, despite the fact that climate change is already affecting the biology and ecology of some organisms, as well as several chemical pathways. Seafood security and safety are related issues because unacceptable standards of food safety that render food unfit for human consumption will also impair seafood security, possibly forcing people to consume seafood that is of lower quality or contaminated, or having higher (bio)availability of chemical contaminants. For chemical

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contaminants, including toxins, a systematic change in marine hydrographical conditions due to a change towards warmer temperatures, reduced salinity and hypoxia may directly affect seafood safety at several levels [53]: (a) increase the input of chemicals contaminants to marine systems and consequently the exposure level, particularly due to flood events; (b) change their chemical forms to more toxic ones and thus the exposure level; (c) increase re-suspension processes of sediment-bound chemical contaminants; (d) increase their bioavailability, especially with regard to metals, with contaminants being converted to more bioavailable forms (e.g. increases in temperature enhance the methylation rate of mercury); (e) diminish the species’ ability to deal with toxic substances and the different physiological regulation processes involved in the detoxification of hazardous substances; and (f) modification of contaminant transport pathways to marine systems. The significance of a particular pathway or process depends on the underlying properties (e.g. hydrophobicity, solubility, volatility) and form of the contaminant/toxin (particulate, particle-associated, dissolved, etc.) [55, 56]. In contrast, climate impacts on plankton can be direct, such as through the effects of temperature and solar radiation on their physiology and growth substrates, or indirect, through caloric and kinetic energy inputs and freshwater inputs, which determine the availability of essential elements, light and reducing power [57]. In any case, predicting the health of Mediterranean marine ecosystems in response to global warming and other anthropogenic phenomena inevitably leads to understanding and predicting the adaptations of the planktonic community to the changing environment [58].

8.2.4 Ecological response to Climate Change Susceptibility of organisms to diseases can be affected by three interlinked factors: the organism itself; contaminants; and the environment. If any of these three factors are altered, changes in the progression of a disease epidemic can occur. Climate change may impact these factors in various ways, such as by exacerbating the presence of biological contaminants in the marine environment (e.g. toxins produced by HABs) and increasing pathogenic microorganisms’ populations [54]. Climate change in its turn encompasses shifts in a number of environmental factors including pH, water level, salinity and temperature [57], and in a number of related changes, e.g. oxygen and food availability, that ultimately modify organism performances and adaptation capability [58]. According to climatic models, the Mediterranean basin is bound to be one of the regions most affected by the ongoing warming trend and by an increase in extreme events. There are reasons to believe that the Mediterranean is already one of the most impacted seas in the world, since climate change interacts synergistically with many other influences [59]. Climate change combines with Atlantic influx, Lessepsian migration and the introduction of exotic

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species by humans to the establishment of tropical marine biota in the Mediterranean Sea [60], which increase the establishment and range extension of tropical exotic species, such as the pufferfish, which contains the strongest neurotoxin known to date, tetrodotoxin [61]. Present-day warming ultimately favors the spread of warmwater species through direct and indirect effects, and especially by changing water circulation [19]. Furthermore, climate change seems to be responsible for a change in the geographical distribution of certain harmful benthic dinoflagellates. Until recently, a number of them were mainly restricted to circumtropical areas, but have lately spread to temperate regions, including the Mediterranean sea. Indeed, records of toxic benthic dinoflagellates have dramatically increased along the Mediterranean coast over the last decades and the list of new/emerging harmful species is continuously growing [24, 62–64]. In particular, Ostreopsis ovata, a species considered as tropical, has bloomed in the Mediterranean region in recent years with increasing frequency, intensity and distribution, in both western and eastern coasts of the Mediterranean, causing benthic organisms’ mortality and human health problems. Health problems also include the formation of toxic aerosols due to wave action, leading to respiratory asthma-like symptoms in humans, such as those reported in the Ostreopsis ovata blooms in the Ligurian coast of Italy in 2005 [65]. In the same context, the tropical genus Gambierdiscus responsible for the production of ciguatera fish poisoning (CFP) ciguatoxins, has been detected in the Mediterranean Sea, as well as in north-eastern Atlantic Ocean, Canary Islands and Madeira [66]. The presence of another recently described benthic dinoflagellate species Vulcanodinium spp., responsible for the production of emerging toxins (pinnatoxins), has been repeatedly reported in the Mediterranean region along with the associated toxins in bivalve mollusks ([67–69]; Greek NRL Marine Biotoxins, unpublished data). Climate change subjects marine ecosystems to multifactorial stressors such as increased temperature, enhanced surface stratification, alteration of ocean currents, intensification or weakening of nutrient upwelling, stimulation of photosynthesis by elevated CO2 , reduced calcification from ocean acidification and changes in land runoff and micronutrient availability. The topic of HABs and climate change is usually addressed by researchers by taking into account only single environmental factors (e.g. CO2 , temperature increase, stratification), single biological properties (photosynthesis, calcification, nutrient uptake) or selected “pet” species categories. Complex factor interactions, on the other hand, are rarely covered by simulated ecophysiological experiments [38]. Nevertheless, it seems there are early signs that some parts of the Mediterranean may become less productive in response to climate-driven increased sea surface temperatures and associated reduced nutrient availability [70]. Ecosystems disturbed by pollution or climate change also tend to be more prone to ballast water invasions [71]. Some harmful algal bloom phenomena, such as toxic dinoflagellates benefiting from land runoff and/or water column stratification as well

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as benthic dinoflagellates responding to coral reef disturbance may increase, while others may diminish in areas currently impacted [65]. Temperature is probably the most widely recognized component of climate change and also plays a crucial role in determining potential algal growth rates. Consequently, temperature can be of significant influence to the community dynamics of HAB species relative to their competitors and grazers. In diatoms, for instance, nitrate uptake and reduction rapidly decline at elevated temperatures, thus potentially favoring competing algae. For instance, the benthic/epiphytic dinoflagellate genus Gambierdiscus spp. respond to warming sea surface temperatures and habitat transformation by concurrent spreading of the marine macroalgae with which they are associated [72–74]. In contrast, where Ostreopsis spp. is concerned, the majority of laboratory experiments examining temperature suggest that Ostreopsis grow more efficiently at high temperatures, but are more toxic at lower temperatures [75–78]. Climatic changes in conjunction with deteriorated ecosystems near ports and lagoons have also resulted in significant changes in biodiversity due to the introduction and establishment of exotic species. The majority of exotics are found in the eastern basin (Levantine) of the Mediterranean Sea. The introduction of exotic species (more than 600 records in 2004) is a dynamic non-stop process with approximately 15 new species reported each year. It is noteworthy that in the 21st century, 64 new species have been reported in the Mediterranean, with 23 of them recorded in 2004 [40].

8.3 Emerging toxins in the Mediterranean Sea Marine biotoxins detected worldwide, but particularly in European waters, were originally classified based on their acute symptomatic effect in humans following intoxication. The three main groups monitored in the European Union (EU), since 1991 by the Directive 91/492/EEC [79], were: (a) Paralytic Shellfish Poisoning (PSP) toxins; (b) Diarrheic Shellfish Poisoning (DSP) toxins; and (c) Amnesic Shellfish Poisoning (ASP). However, due to the progress of alternative detection methods, classification of DSP toxins has changed to focus more on the chemical structures and properties of the toxins. DSP toxins have therefore become known as lipophilic toxins incorporating okadaic acid, dinophysistoxins, azaspiracids, pectenotoxins and yessotoxins, with the last two not proved to cause diarrheic symptoms following intoxication. For each of these three main toxin groups and subgroups, the occurrence of the toxins, their chemical characteristics, toxicokinetic evaluations, human-exposure assessments and detailed review of potential methods of analysis have in recent years been published by the European Food Safety Authority (EFSA) as scientific opinions [80–85]. The diversity of the numerous analogues or natural enzymatic metabolites of marine biotoxins has been described [86]. However, this series of reviews also included prospective emerging toxins to European waters such as cyclic imines, palytoxin, tetrodotoxin, maitotoxin, ciguatoxins and the neurotoxin-poisoning brevetoxins, as

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their occurrence could have severe implications with regard to seafood safety [87]. Although EFSA has published the relevant scientific opinions for emerging toxins [88–91], with the exception of tetrodotoxin, these toxins are not yet specified by the current EU regulations [92].

8.3.1 Identified emerging toxins and climate change effects 8.3.1.1 Tetrodotoxin Tetrodotoxin (TTX) is one of the most potent low-molecular-weight marine neurotoxins (319 Da). Its chemical structure was described as a cage-like polar molecule with a cyclic guanidinium moiety fused to a dioxy-adamantane skeleton embellished by six hydroxyl groups. To date, 29 different analogues of TTX have been reported. Their degree of toxicity varies among analogues, although not much is known about them [93]. An important recognized feature is that the deoxy analogues of TTX are less toxic than TTX, while the hydroxyl analogues are more toxic than TTX [94]. TTX is considered to be produced by certain endo-symbiotic bacteria, such as Vibrio sp., Pseudomonas sp. and Alteromonas [95]. Similarly to saxitoxin (STX), TTX consumption induces severe symptomatology in humans starting from numbness and mild gastrointestinal effects to respiratory paralysis and even death of human consumers. Consequently, bioaccumulation of TTX in seafood and subsequent entrance in the human food chain poses a real and very significant risk to human safety. TTX intoxication is most commonly associated with the consumption of puffer fish and sometimes by the ingestion of gastropods or crabs. The close relationship between TTX and the tropical environment, where for a long time most of the species bearing this toxin lived, explains the fact that until recently the large majority of the reported poisoning cases from this toxin were confined to the southwest Asian area, especially in Japan were the regular consumption of Fugu-related cuisine results in most of the registered events. Despite this initial geographical limitation, a visible increase of TTX intoxication cases in Mediterranean waters where such cases should be unlikely is occurring [96]. In the last decade, TTX has been found in marine organisms collected from European countries and specifically pufferfish [33, 97, 98] the marine gastropod Charonia lampas lampas [97–99] and very recently in cultured mussels of the species Mytilus galloprovincialis in Greece [Vlamis et al., under preparation]. This new phenomenon of displaced TTX detection cases and the global increase of water temperatures can be linked to both the increase of the TTX-vector’s presence in Mediterranean waters, as well as to the increase of TTX contents in the actual vectors. A number of researchers attribute the new occurrence of TTX in European regions to the so-called “Lessepsian migration” (Fig. 8.1). In 1869, the opening of Suez Canal caused the migration of many Red Sea species through the new waterway, which have settled and have been well established in the Eastern Mediterranean. These migrations are a growing phenomenon that closely accompany the increase in global tem-

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perature [100]. Among the most significant migrations in the Mediterranean Sea is that of fishes from the Lagocephalus family, known vectors for TTX, such as Lagocephalus sceleratus [30, 33, 101, 102], L. suezensis [103] and L. lagocephalus [104]. Concerning the case of L. sceleratus, one of the most recent Lessepsian species discovered, it has been related to cases of intoxication [102] along the Mediterranean shore and its presence in the artisanal fishing captures is an increasing observation [96]. Published studies are rather inconsistent when relating a water temperature increase to the increase of TTX content in its vectors, although certain studies supporting this argument exist. Matsumoto et al. [105] clearly relate an increased intake rate of TTX into the liver tissue of Takifugu rubripes with the increase in environmental temperature; however, this relation between temperature variation and TTX toxicity potential remains obscure as is the case with the mechanisms involving excretion and accumulation of the toxin on puffer fish [96]. The seasonal rise in environmental temperatures is considered to be a determining factor in the toxicity variation of TTX vectors. Several such studies have attempted to establish a valid correlation between the two. For instance, tests performed in specimens of Taiwanese gobies were performed and showed that a small percentage of the animals had a measurable TTX content between August 1996 to July 1998; this was observed as both regional and seasonal variations of the toxin, with higher values from March to November [106]. These test results differ with those from another study performed during an annual period (August 2000 to August 2001) in Indonesia on the puffer fish Lagocephalus lunaris, which showed that the animals remained toxic through 9 months (March to November), with the toxicity peak being observed in August with 100 % of the test animals possessing the toxin [96, 107]. Still there is an obvious lack of systematic studies and more research is required in order to establish any kind of definitive hypothesis on the seasonal variance of TTX vector toxicity, also taking into account that higher TTX contents are also associated with the spawning period of pufferfish, a period which coincides with the summer months in the Mediterranean Sea [108, 109]. Despite the fact that the epidemiologic perspective of TTX is well studied, its molecular and cellular mechanisms are still not clear enough. As indicated above, certain bacteria of the microflora are thought to be responsible for TTX production, but no studies on their kinetics are available to understand their possible interactions with the current changing climate patterns. One of the scarce studies to relate changes in water temperature with changes in the bacterial content of TTX vectors (in this case the puffer fish Fugu niphobles) demonstrated that the bacterial content of the skin, gills and intestines of the fish was actually affected by temperature variations. Specifically, the identification of bacteria of the genus Vibrio, known TTX-producers, was positive in temperatures of 20 °C and 29 °C, but negative at 10 °C. These results were further confirmed in laboratory conditions when the same bacteria were cultivated, when again all strains were able to grow at 20 °C and 29 °C, but very few were able to do so at 10 °C suggesting their preference for higher water temperatures [96, 110].

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Ballast waters on the other hand can also be responsible for the transfer of TTXcontaining organisms from Asian waters to European waters. Over the last 20 years, spreading of marine mucilage in the Mediterranean Sea has also been observed due to sea surface warming, which helps the survival of migrated species in this marine environment [94, 111]. The Japanese government has long before established a regulatory limit of 2 mg/kg of TTX equivalents in food due to “fugu” consumption. In contrast, no regulatory limits have been set in the European Union because TTX poisoning had not ever been considered as a potential problem. However, in October 2007, the first toxic European episode was reported in Malaga (Spain), caused by the ingestion of a trumpet shell of the species Charonia lampas lampas. The product was purchased in a Malaga market, but was originally caught in the south coast of Portugal [112]. Again, in 2005 and 2008, pufferfish L. sceleratus present in the Mediterranean Sea due to the Lessepsian migration phenomenon have been incriminated for several severe poisonings and deaths (respiratory distress a few minutes after the meal by sudden paralysis due to the brutal decrease of neuromuscular transmission) following consumption of this highly toxic invader in Israel and Lebanon [102, 113]. This fish is also largely present in Turkey and Greece [33, 104], whereas despite its rather recent introduction, L. sceleratus has already expanded in the central part of the Mediterranean, with proven presence in Tunisia [114], Libya [115] and lately in Italy [116], while its arrival in the western Mediterranean seems to be imminent [117]. An equally toxic indigenous pufferfish species Lagocephalus lagocephalus is also sporadically observed along the North African Mediterranean coast from Morocco to Libya, but this pelagic species from the Atlantic is rarely caught [118]. It is therefore evident that Mediterranean seafood is endangered by being contaminated with this hazardous toxin, which highlights the need to have a regulation and to develop unambiguous, fast and reliable methods to specifically detect and quantify TTX in order to protect human health.

8.3.1.2 Palytoxin Palytoxins (PLTXs) have been originally detected in marine zoanthids (soft corals) of the genus Palythoa, but thereafter the production of several analogs was also confirmed in benthic dinoflagellates of the genus Ostreopsis (e.g. Ostreopsis siamensis, O. mascarenensis, O. ovata). PLTXs were first reported in Hawaii and Japan, in warm waters were the soft corals naturally occur, but are currently known to be distributed worldwide [119, 120], especially after the uncovering of the broad distribution of Ostreopsis spp. Actually in the past decade, blooms of Ostreopsis spp. are increasingly being reported in European countries [121, 122], such as in France, Greece, Italy and Spain, and PLTX-like compounds have been identified in the Mediterranean strains of O. cf. ovata, such as putative palytoxin (pPLTX) [122], ovatoxins-a [122], -b, -c, -d [123] and -f [124]. Recently, a bloom of Ostreopsis spp. on the coast of Algarve (south

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of Portugal) showed that species capable of producing palytoxin analogues such as ovatoxin may be spreading from the Mediterranean to the north Atlantic [125, 126]. Temperature is a crucial factor determining both growth potential as well as toxin production of the genus Ostreopsis. Clonal laboratory cultures of O. lenticularis exposed to elevated temperatures (30–31 °C) for 33 and 54 days showed significant increases in the quantity of extractable toxin they produced as compared to their toxicities versus cells grown at temperatures of 25–26 °C. Furthermore, O. lenticularis samples collected directly from the field following exposure to elevated temperatures for comparable periods of time also showed significant increases in extractable toxin [127]. The increased toxicity of both field sampled and laboratory grown O. lenticularis exposed to elevated temperatures is considered to result from the effects of elevated temperatures on their metabolism and/or the symbiotic bacterial found associated with these microalgae. The number of bacteria associated with cultured O. lenticularis exposed to elevated temperatures was significantly reduced. Increased toxin recovery from O. lenticularis exposed to elevated temperatures, on the other hand, may have resulted from the direct effect of temperature on toxin production and/or the reduction of Ostreopsis associated bacterial flora that consume toxin in the process of their growth. This reduction in the quantity of associated bacterial flora in temperature treated cultures, in its turn, may result in increased toxin recovery from O. lenticularis due to a reduction in the consumption of toxin by these symbiotic bacteria [127]. On the other hand, the optimum temperatures for growth and toxicity of O. ovata were found to be inversely related. High water temperatures (26–30 °C) stimulated O. ovata cells growth rate and biomass accumulation and low toxicities while lower temperatures (20–22 °C) induced higher toxicity per cell and lower cell numbers. Based on these results, it was suggested that increased sea surface temperature, which can result from global warming, may play a crucial role in inducing the geographical expansion and biomass increase in blooms of O. ovata in the future [77]. In another study, it was also concluded that environmental conditions seem to play a key role in influencing the abundance of Ostreopsis spp. High cell densities of an Adriatic O. cf. ovata isolate were generally recorded in concomitance with relatively high temperature, salinity and low hydrodynamic conditions. The highest growth rates of that Adriatic strain were recorded for cultures grown at 20 °C and at salinity values of 36 and 40, in accordance with natural bloom surveys. Toxicity was also affected by growth conditions, with the highest toxin content on a per cell basis being measured at 25 °C and salinity 32. However, the highest total toxin content on a per liter basis was recorded at 20 °C and salinity 36, since under such conditions the growth yield was the highest [128]. Similarly, O. cf. ovata from Japanese waters was found to grow faster than other benthic toxic-dinoflagellates [Coolia monotis, Gambierdiscus toxicus and Prorocentrum lima (Dinophyceae)] in waters of high temperature and salinity. This physiological feature was considered to confer an ecological advantage on O. cf. ovata in the bloom development during warmer seasons and could be responsible for outbreaks of PTX-like poisoning, especially during the warmer seasons [129].

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The palytoxin group (PLTXs) constitutes one of the most poisonous non-protein marine toxins known to date. They present high acute toxicity in animals by the intravenous or intraperitoneal route (e.g. an i.v. lethal dose (LD50 ) ranging from 0.15 to 0.73 μg/kg in mice, but the oral route has been reported as the least sensitive [130]. Despite this fact, a number of acute toxicity cases and deaths have been reported from human outbreaks; but reliable quantitative data on acute toxicity in humans are still unavailable. In view of the acute toxicity reports and the lack of chronic toxicity data for the PLTX-group toxins, the EFSA Panel on Contaminants was able to derive an oral acute reference dose (ARfD) of only 0.2 μg/kg b.w. for the sum of PLTX and its analog ostreocin-D. In order for a 60 kg adult to avoid exceeding the ARfD, a 400 g portion of shellfish meat should not contain more than 12 μg of the sum of PLTX and ostreocin-D, corresponding to 30 μg/kg shellfish meat [88]. Despite the fact that there are increasing records of the PLTXs presence largely above these levels in many edible marine organisms from the Mediterranean Sea, the European Union has not yet adopted a maximum permissible limit to confront the risk of PLTX poisoning of European consumers [9, 131].

8.3.1.3 Cyclic imines (Gymnodimine, Spirolides, Pinnatoxins) The emerging toxin group of cyclic imines (CIs) consists of spirolides (SPXs), gymnodimines (GYMs), pinnatoxins (PnTXs) and pteriatoxins (PtTXs) and is a family of marine biotoxins largely present in shellfish and other marine organisms. They are macrocyclic compounds with imine (carbon-nitrogen double bond) and spiro-linked ether moieties. They have been grouped together because of their common imine group as a part of a cyclic ring, which confers the pharmacological and toxicological activity, and due to their similar acute “fast acting toxicity” in mice [90, 132]. In addition to SPXs, GYMs, PnTXs and PtTXs, the CI group comprises prorocentrolides and spiroprorocentrimines, which to date have not been reported in European shellfish. Although SPXs, GYMs, PnTXs and PtTXs are now known to occur in microalgae and/or shellfish in several parts of the world (Canada, Denmark, New Zealand, Norway, Scotland, Tunisia, USA and Japan), no information has been reported linking these toxin groups to poisoning events in humans [133–136]. This fact explains the absence of regulatory limits and official analysis methods for CIs in shellfish. The relevant EFSA opinion also concluded that, at the time it was issued, estimated exposure to SPXs did not raise concern for the health of the consumer, although it was stressed that this conclusion for SPXs was based on very limited toxicity data [90]. However, the working group of the EU-RLMB has proposed a guidance level for the sum of SPXs of 400 μg/kg in shellfish meat [137, 138]. CIs are produced by different dinoflagellates to a different extent: SPXs are mainly produced by Alexandrium ostenfeldii also known as A. peruvianum [139, 140], GYMs are produced by Karenia selliformis, also known as Gymnodinium selliforme [141]. The PnTXs-producing organism has been described as a peridinoid dinoflagellate [135]

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and recently identified as Vulcanodinium rugosum [67], prorocentrolides have been isolated from Prorocentrum lima [142], spiro-prorocentrimines are suggested to be produced by Prorocentrum species [143] and PtTXs are suggested to be bio-transformation products of PnTXs in shellfish, so no producing organism has been identified [90]. SPXs have been identified in a number of European countries bordering the Mediterranean Sea, Atlantic coast and North Sea. SPXs have been found in their producer dinoflagellate A. ostenfeldii/peruvianum in Scotland [144], Italy [145], Denmark [146] and Ireland [140]. They have also been found in shellfish from Norway [147], Spain [148] and in Italy [138]. SPXs are also commonly detected in Greek shellfish since 2008 [149], at concentrations ranging from trace levels up to 118 μg/kg, with the highest levels producing positive MBA assays (Greek NRL Marine Biotoxins, unpublished data). GYMs have been first found in shellfish from New Zealand, specifically in greenshell mussel, blue mussel, scallop, cockle, surf clam, oyster and abalone [150, 151]. To date, there are no published records for the presence of GYMs in shellfish produced in Europe, but they have been reported in products imported from outside of Europe. We have, however, repeatedly determined GYMs in Greek shellfish (mussels, venus clams and hard clams) at concentrations ranging from trace levels up to 66 μg/kg within the framework of the Greek HAB monitoring program. The highest concentrations coincided also with positive MBA tests (Greek NRL Marine Biotoxins, unpublished data). With regard to other Mediterranean countries, GYM has also been identified in clams Ruditapes decussatus from the Gulf of Gabes in Tunisia [152–154]. The low acute toxicity of gymnodimine when ingested with food (> 7500 μg/kg in mice) suggests that this compound is of low risk to humans, a conclusion consonant with anecdotal evidence for the absence of harmful effects in individuals who had consumed shellfish contaminated with gymnodimine [155]. Despite this fact, the chronic toxicity of GYM remains unclear as its role in the development of neurodegenerative illnesses like Alzheimer or Parkinson’s diseases has been debated [156, 157]. PnTXs were quite recently identified for the first time in shellfish in Europe. They have so far been found in Norwegian mussels [158], French mussels and clams [68] and Spanish mussels and oysters [69]. PnTX-G is also commonly detected in Greek shellfish, with the highest concentrations occurring in summer months, whereas there is also evidence for the presence of the relevant fatty ester metabolites (Greek NRL Marine Biotoxins, unpublished data). The recent findings suggest that PtTXs are transformed from PnTXs in shellfish, while their presence in Europe has only been reported so far at non-quantifiable traces in French shellfish [68]. Alexandrium ostenfeldii has been widely observed in temperate waters of Europe [159], North America [139], the Russian Arctic [160] and Eastern Siberian Seas [161]. There are also records of the occurrence of A. ostenfeldii from the coast of Spain [162], the Mediterranean [163], New Zealand [164], Peru [165] and Japan [166]. However, for a long time, A. ostenfeldii has been considered mainly as a background species, occurring at low cell concentrations mixed with other bloom forming dinoflagellates

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[144, 159, 167]. Only in the past decade has it gained increasing attention when dense blooms of this species (or its synonym A. peruvianum) were reported, e.g. from South America [165], the Northern Baltic Sea [168], along the Adriatic coast of Italy [145], the estuaries of the US East coast [169] and, most recently, the Netherlands [170]. It is not clear whether the recent increase in bloom events is due to anthropogenic spreading or changing environmental conditions favoring bloom formation. Most of the recent blooms occurred during summer in coastal areas and were associated with warm water periods (e.g. [171]). Experimental studies indicate that increased water temperature has a favorable effect on A. ostenfeldii bloom populations and it has been suggested that changing climate conditions promote bloom formation [172]. The species produces PSP toxins [173], spirolides [139] and gymnodimines [174], and all compounds may even occur together in one strain [169]. Thus, an increase of A. ostenfeldii bloom events with several potent toxins involved may represent a new risk to the environment that is associated with climate change [175]. The situation is further complicated by the fact that toxin composition of A. ostenfeldii was consistently altered by elevated temperature and increased CO2 supply in the strains tested by Kremp et al. [172], resulting in an overall promotion of saxitoxin production, with potentially severe consequences for the coastal ecosystem. In a study regarding growth and toxicity of a cultured strain of Karenia seliformis, where different temperature and salinity combinations were tested, it was concluded that growth rates were similar for the different salinities tested, but showed an increasing trend with water temperature increasing from 15–17 °C to 20–21 °C. Toxicity of the culture on the other hand seemed to be dependent on culture age and growth phase, with the declining phase coinciding with higher GYM production [176]. Vulcanodinium rugosum [67] was first described from water samples of Mediterranean lagoons, with a water temperature of 23.3 °C and a salinity of 36.5 psu. In another study, V. rugosum cells were isolated and cultured from the tropical Mexican Pacific, where the local conditions of surface temperature and salinity, in which the cells were found, were 28 °C and 35 psu, respectively, and the culture conditions were similar. It was concluded that the species seemed to prefer warm waters and fully marine environments [177]. Vulcanodinium spp. is now also quite common in Greek waters, especially during the warm periods of the year (unpublished regulatory monitoring data, Greek Ministry of Rural Development and Food). The fact that, to date, this genus has been detected in Australia, New Zealand, Japan [178], Mexico [177] and coastal Mediterranean areas (France, Greece) is also indicative of the species’ preference to warmer temperatures, thus the temperature increase resulting from climate change could create risks by providing more optimal conditions for its growth. Currently the situation remains that there are no regulations on CIs in shellfish in Europe or in other regions of the world, as no acute human intoxication by consumption of shellfish contaminated by PnTXs has been reported. Given the novelty of the research area of CIs, it is crucial to better describe their mechanisms of action, as well as to widen the toxicological and pharmacological data in order to determine if they

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pose a public health risk. This is particularly important considering the widespread distribution of CIs in seafood and their potent binding to nicotinic acetylcholine receptors in the central and peripheral nervous systems; this might have long-term effects on human health. It would also be important to develop suitable methodologies for the detection of all CIs and implement preventive measures in monitoring programs [179].

8.3.1.4 Ciguatoxins Ciguatoxin (CTX)-group toxins are marine biotoxins which occur in fish as a result of biotransformation of the precursor gambiertoxins produced by the benthic dinoflagellate Gambierdiscus toxicus [180–183]. They are mainly found in Pacific, Caribbean and Indian Ocean regions and they are classified as Pacific (P), Caribbean (C) and Indian Ocean (I) CTX-group toxins. Recently, however, Gambierdiscus sp. has been also detected in the Mediterranean Sea, whereas CTX-group toxins were identified for the first time in fish caught in Europe. The first report in Europe on the occurrence of Gambierdiscus sp., potential producer of ciguatoxin (CTX), was in Crete Island, Greece in 2003 [184]. Later, the intoxication of fishermen who ate fish caught off the Madeira island archipelago revealed the presence of CTXs [185]. So far, no regulatory limits exist for CTXs in fish in Europe, but the legislation requires that no fish products containing CTXs be placed on the market [89]. A Guidance Level of 40 ng ciguatoxin/kg fish has been recommended by the working group of the EU-RLMB [137]. In contrast, the United States Food and Drug Administration (US FDA) has proposed guidance levels of ≤ 0.1 μg/kg of C-CTX-1 equivalents and ≤ 0.01 μg/kg of P-CTX-1 (initially named CTX-1B, both names currently in use) equivalents [186]. In addition, since the unique existing certified standard is for P-CTX-1, toxicity equivalence factors (TEFs) for CTX congeners have been established by acute toxicity in mice (LD50 ) as follows: P-CTX-1 = 1, P-CTX-2 = 0.3, P-CTX-3 = 0.3, P-CTX-3C = 0.2, 2,3-dihydroxy P-CTX-3C = 0.1, 51-hydroxy P-CTX-3C = 1, P-CTX-4A = 0.1, P-CTX-4B = 0.05, C-CTX-1 = 0.1 and C-CTX-2 = 0.3. These TEFs should be applied to express individual analogues identified with quantitative detection methods as P-CTX-1 equivalents [89, 132]. CTXs can enter into the food webs through the consumption of dinoflagellates by herbivorous fish and their subsequent consumption by carnivorous fish [187]. Humans would be intoxicated by consumption of herbivorous and carnivorous fish containing CTXs. CTX precursors accumulate in fish tissue (mainly in viscera, but also in the muscle or other parts [188]) and may afterwards be metabolized into different CTX forms, which are responsible for human intoxication. Around 30 analogues from the Pacific Ocean (P-CTXs) [180–182, 189–192], from the Caribbean Sea (C-CTXs) [193] and from the Indian Ocean (I-CTXs) [194] have been identified. Ciguatera fish poisoning (CFP) is the foodborne illness is responsible for the highest reported incidence of human poisoning from seafood consumption worldwide [66, 182]. At present, an estimation indicates that between 10 000 and 50 000 people suffer around the world from this disease annually [89]. CFP is endemically found in Indo-

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Pacific and Caribbean areas. However, in recent years, CTXs are appearing in countries not expected for their latitude, such as waters close to European and African continents, e.g. in the Canary Islands (Spain) [195] and Madeira (Portugal) [185, 196]. The presence of ciguatoxin-like substances in edible fish of the eastern Mediterranean of Israel was also suspected in the case of an assumed ciguatera poisoning incident in this atypical site after consumption of the fish Siganus rivulatus [197]. This northern expansion has been attributed, for instance, to changes on the distribution of toxinproducing microalgae to northern latitudes [198] or to migration of fish containing the toxins [199]. In support of the former, the presence of Gambierdiscus sp. has been recorded in the eastern Mediterranean Sea, from 2003 in Crete Island [184]. Prior to this recording, the northernmost area where it had been found was the Canary Islands – approximately 28° N [184, 200, 201]. The aforementioned finding in Crete and later on the detection of Gambierdiscus in Cyprus increased the latitude to which this genus is distributed from 28° N to just above 35° N. However, the northern geographical boundaries of Gambierdiscus distribution were further expanded by the detection of Gambierdiscus on February 2009 in Saronicos Gulf, Salamina Island, at a latitude of about 38° N. This record, situated in the middle latitude of the Mediterranean Sea, constitutes the current northernmost point of Gambierdiscus distribution worldwide [63]. All these findings, thus, suggest a possible future concern about CTX-group toxins in fish and seafood originating from Europe. The occurrence of representatives of what was once considered a tropical or subtropical genus, like Gambierdiscus, is in accordance with the suggestions of several researchers regarding climate change impact on the geographical expansion of tropical microalgae, and the “tropicalization” of the Mediterranean Sea in recent years. This fact, together with the Ostreopsis species range expansion in the Mediterranean during the last decade, and their toxicity, could possibly constitute a serious threat to human health by both ciguatera and palytoxin intoxications [201].

8.3.1.5 Brevetoxin Brevetoxin (BTX) group toxins cause Neurotoxic Shellfish Poisoning (NSP), a syndrome characterized by mainly neurological and gastrointestinal effects. Symptoms and signs of NSP include e.g. nausea, vomiting, diarrhoea, parasthesia, cramps, bronchoconstriction, paralysis, seizures and coma. They typically occur within 30 minutes to 3 hours of consuming contaminated shellfish and last for a few days. Persistent symptoms and fatalities have not been reported. Dermal or inhalation exposure can result in irritant effects. BTX-group toxins are metabolized in shellfish and fish, and several metabolites of BTX-group toxins have been characterized. Consumers of contaminated shellfish and fish are thus primarily exposed to BTX-group toxin metabolites rather than parent algal BTX-group toxins.

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BTXs are marine biotoxins that can accumulate in shellfish and fish, and are primarily produced by the dinoflagellate Karenia brevis (formerly called Gymnodinium breve and Ptychodiscus brevis). So far, NSP appears to be limited to the Gulf of Mexico, the east coast of the United States of America and the New Zealand Hauraki Gulf region. To date BTX-group toxins have not been reported in shellfish or fish from Europe. However, the discovery of new BTX-group toxin producing algae such as Chattonella antiqua, Chattonella marina, Fibrocapsa japonica, Heterosigma akashiwo, which have been reported to produce BTX-like toxins [202], and the apparent trend towards expansion of algal bloom distribution suggest that BTX-group toxins are emerging in other regions in the world [91, 133]. Karenia brevis (Gymnodinium breve) populations are found in warm temperate to tropical waters, most regularly from the Gulf of Mexico, off the west coast of Florida. However, K. brevis and K. brevis-like species have also been reported from the West Atlantic, Spain, Greece, Japan and New Zealand [203–205]. Blooms initiate offshore requiring high salinities (> 30 ‰) and high temperatures [205–208]. K. brevis grows at a temperature range between 4 and 33 °C; however, its optimal growth range is 22–28 °C. In addition, this organism can live in a salinity of between 25–45 psu, with an optimum of 30–34 psu [209]. Similarly, ocean conditions that could arise from climate change, such as lowered pH and/or increased temperature, were shown to promote significant increases in growth rates of Heterosigma akashiwo compared to controls [210–212]. Since the summer of 1994, the Raphidophyceae Fibrocapsa japonica has become the most recurrent blooming species in the Western Adriatic coastal area causing heavy water discoloration, with cells concentrations exceeding 106 cells/l. During these blooms, other Raphidophyceae (Heterosigma akashiwo, Chattonella spp.) were usually present in very low concentrations; however, these species can exceptionally form nearly monospecific blooms as occurred in July 2011 for H. akashiwo and C. globosa [213]. H. akashiwo has been also reported in other sites of the Mediterranean Sea, e.g. in Turkey (Izmir Bay) [214, 215], Egypt [216] and Syria (Lattakia port) [217]. It is thus possible that climate change could trigger blooms and/or toxin production from these species in the future. Currently there are no regulations on BTX-group toxins in shellfish or fish in Europe; the EFSA CONTAM Panel could not comment on the risk associated with the BTX-group toxins in shellfish and fish that could reach the European market due to the lack of occurrence data on shellfish or fish in Europe, the limited data on acute toxicity and the lack of data on chronic toxicity [91]. However, some countries in other regions of the world have set action levels or maximum levels for BTX-group toxins in shellfish. In the USA, the action level is 20 mouse units (MUs)/100 g (0.8 mg BTX-2 equivalents/kg shellfish) [218]. In New Zealand and Australia the maximum level for BTX-group toxins is 20 MUs/100 g, but the BTX analogue is not specified [91, 219, 220].

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8.4 Conclusion Climate change has become one of the most critical issues for the future of our planet. It involves significant changes in the variability or average state of the atmosphere, such as related to temperature, precipitation and/or wind patterns, over durations ranging from decades to millions of years. Climate change effects will have implications for food production, food security and food safety. In particular, the safety of feed and food products arising from marine production systems is expected to be affected by climate change, specifically by increased occurrence of phycotoxins (marine biotoxins). The apparent increase in the occurrence of HABs and the recognition that climate changes may be creating a marine environment particularly suited to HAB-forming species of algae highlight the need for governments to ensure that existing risk management measures are sufficient and are in accordance with international recommendations. Countries are encouraged to implement integrated shellfish and micro-algal monitoring programs, especially for emerging toxins, as part of Marine Biotoxin Management Plans to strengthen risk management capability and to enhance consumer protection.

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[167] Moestrup Ø, Hansen, PJ. On the occurrence of the potentially toxic dinoflagellates Alexandrium tamarense (= Gonyaulax excavata) and A. ostenfeldii in Danish and Faroese waters. Ophelia 1988;28:195–213. [168] Kremp A, Lindholm T, Dreßler N, Erler K, Gerds G, Eirtovaara S, Leskinen E. Bloom forming Alexandrium ostenfeldii (Dinophyceae) in shallow waters of the Aland Archipelago, Northern Baltic Sea. Harmful Algae 2009;8:318–28. [169] Tomas CR, Van Wagoner RM Tatters AO, White KD, Hall S, Wright JLC. Alexandrium peruvianum (Balech and Mediola) Balech and Tangen a new toxic species for coastal North Carolina. Harmful Algae 2012;17:54–63. [170] Burson A, Matthijs HCP, de Bruijne W, Talens R, Hoogenboom R, Gerssen A, Visser PM, Stomp M, Steur K, van Scheppingen Y, Huisman J. Termination of a toxic Alexandrium bloom with hydrogen peroxide. Harmful Algae 2014;31:125–35. [171] Hakanen P, Suikkanen S, Franzén J, Franzén, H, Kankaanpää H, Kremp A. Bloom and toxin dynamics of Alexandrium ostenfeldii in a shallow embayment at the SW coast of Finland, northern Baltic Sea. Harmful Algae 2012;15:91–9. [172] Kremp A, Godhe A, Egardt J, Dupont S, Suikkanen S, Casabianca S, Penna A. Intraspecific variability in the response of bloom forming marine microalgae to changing climatic conditions. Ecol Evol 2012;2:1195–207. [173] Hansen PJ, Cembella AD, Moestrup Ø. The marine dinoflagellate Alexandrium ostenfeldii: paralytic shellfish toxin concentration, composition, and toxicity to a tintinnid ciliate. J Phycol 1992;28:597–603. [174] Van Wagoner RM, Misner I, Tomas CR, Wright JLC. Occurrence of 12-methylgymnodimine in a spirolide-producing dinoflagellate Alexandrium peruvianum and the biogenetic implications. Tetrahedron Lett 2011;52:4243–6. [175] Tillmann U, Kremp A, Tahvanainen P, Krock B. Characterization of spirolide producing Alexandrium ostenfeldii (Dinophyceae) from the western Arctic. Harmful Algae 2014;39:259–70. [176] Medhioub A, Medhioub W, Amzil Z, Sibat M, Bardouil M, Ben Neila I, Mezghani S, Hamza A, Lassus P. Influence of environmental parameters on Karenia selliformis toxin content in culture. Cah Biol Mar 2009;50:333–42 [177] Hernández-Becerril DU, Rodríguez-Palacio MC, Lozano-Ramírez C. Morphology and life stages of the potentially pinnatoxin-producing thecate dinoflagellate Vulcanodinium rugosum from the tropical Mexican Pacific. Botanica Marina 2013;56:535–40. [178] Rhodes L, Smith K, Selwood A, McNabb P, Munday R, Suda S, Molenaar S, Hallegraeff G. Dinoflagellate Vulcanodinium rugosum identified as the causative organism of pinnatoxins in Australia, New Zealand and Japan. Phycologia 2011;50:624–8. [179] Stivala CE, Benoit E, Araoz R, Servent D, Novikov A, Molgo J, Zakarian A. Synthesis and biology of cyclic imine toxins, an emerging class of potent, globally distributed marine toxins. Nat Prod Rep 2014: doi:10.1039/C4NP00089G. [180] Murata M, Legrand AM, Ishibashi Y, Yasumoto T. Structures of ciguatoxin and its congener. J Am Chem Soc 1989;111:8929–31. [181] Murata M, Legrand AN, Ishibashi Y, Fukui M, Yasumoto T. Conformations of ciguatoxin and related polyethers. Abstr Pap Am Chem Soc 1990;200:54-AGFD. [182] Lehane L, Lewis RJ. Ciguatera: Recent advances but the risk remains. Int J Food Microbiol 2000;61:91–125. [183] Lehane L. Ciguatera update. Medical Journal of Australia 2000;172:176–179. [184] Aligizaki K, Nikolaidis G. Morphological identification of two tropical dinoflagellates of the genera Gambierdiscus and Sinophysis in the Mediterranean Sea J Biol Res-Thessalon 2008; 9:75–82.

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[185] Otero P, Pérez S, Alfonso A, Vale C, Rodríguez P, Gouveia NN, Gouveia N, Delgado J, Vale P, Hirama M, Ishihara Y, Molgó J, Botana LM. First toxin profile of ciguateric fish in Madeira Arquipelago (Europe). Anal Chem 201;82:6032–9. [186] CDC (Centers for Disease Control and Prevention). Cluster of Ciguatera Fish Poisoning, 2007. Morbidity and Mortality Weekly Report (MMWR). North Carolina, NC, USA; CDC: 2009. p. 283–5. [187] Mills AR. Poisonous fish in the South Pacific. J Trop Med Hyg 1956;59:99–103. [188] Vernoux JP, Lahlou N, Elandaloussi SA, Riyeche N, Magras LP. A study of the distribution of ciguatoxin in indiviual Caribbean fish. Acta Trop 1985;42:225–233. [189] Lewis RJ, Sellin M, Poli MA, Norton RS, MacLeod JK, Sheil MM. Purification and characterization of ciguatoxins from moray eel (Lycodontis javanicus, Muraenidae). Toxicon 1991;29: 1115–27. [190] Satake M, Ishimaru T, Legrand AM, Yasumoto T. Isolation of a ciguatoxin analog from cultures of Gambierdiscus toxicus. In: Samyda TJ, Shimizu Y. [ed.] Toxic Pyhtoplankton Blooms in the Sea. New York, NY, USA; Elsevier; 1993. p. 575–9. [191] Satake M, Murata M, Yasumoto T. The structure of CTX3C, a ciguatoxin congener isolated from cultured Gambierdiscus toxicus. Tetrahedron Lett 1993;34:1975–8. [192] Satake M, Fukui M, Legrand AM, Cruchet P, Yasumoto T. Isolation and structures of new ciguatoxin analogs, 2,3-dihydroxyCTX3C and 51-hydroxyCTX3C, accumulated in tropical reef fish. Tetrahedron Lett 1998;39:1197–8. [193] Lewis RJ, Vernoux JP, Brereton IM. Structure of Caribbean ciguatoxin isolated from Caranx latus. J Am Chem Soc 1998;120:5914–20. [194] Hamilton B, Hurbungs M, Vernoux JP, Jones A, Lewis RJ. Isolation and characterisation of Indian Ocean ciguatoxin. Toxicon 2002;40:685–93. [195] Perez-Arellano JL, Luzardo OP, Brito AP, Cabrera MH, Zumbado M, Carranza C, Angel-Moreno A, Dickey RW, Boada LD. Ciguatera fish poisoning, Canary Islands. Emerg Infect Dis 2005; 11:1981–2. [196] Gouveia N, Delgado J, Vale P. Primeiro registo da ocorrência de episódios do tipo ciguatérico no arquipélago da Madeira. In Abstract Book of X Reuniao Oberica, Fitoplancton Toxico e Biotoxinas. Lisbon, Portugal; IPIMAR: 2009. [197] Bentur Y, Spanier E. Ciguatoxin-like substances in edible fish on the eastern Mediterranean. Clin Toxicol 2007;45:695–700. [198] Fraga S. Global climate change and harmful algal blooms (HABs). In: Abstract Book of 4th European Phycological Congress. Oviedo, Spain; Elsevier Science: 2007. p. 41. [199] Stebbing ARD, Turk SMT, Wheeler A, Clarke KR. Immigration of southern fish species to south-west England linked to warming of the North Atlantic (1960–2001). J Mar Biol Assoc UK 2002;82:177–180. [200] Fraga S, Riobó P, Diogène J, Paz, B, Franco JM. Toxic and potentially toxic benthic dinoflagellates observed in Macaronesia (NE Atlantic Archipelagos). In: Programme and Abstracts. XI Int Conf Harmful Algae: 14-19 November, Capetown 2004. p. 115. [201] Aligizaki K, Nikolaidis G, Fraga S. Is Gambierdiscus expanding to new areas? Harmful Algae News 2008;36:6–7. [202] FAO (Food and Agriculture Organization of the United Nations). Marine biotoxins. FAO Food and nutrition paper 2004;80:1–287. [203] Fukuyo Y, Takano H, Chihara M, Matsuoka K. Red Tide Organisms in Japan. An Illustrated Taxonomic Guide. Tokyo; Uchida Rokakuho, Co. Ltd.: 1990. p. 407. [204] Taylor FJR, Fukuyo Y, Larsen J. Taxonomy of harmful dinoflagellates. In: Hallegraeff GM, Anderson DM Cembella AD. [ed.] Manual on Harmful Marine Microalgae. France; IOC Manuals and Guides No. 33. UNESCO: 1995. p. 283–317.

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[205] Steidinger KA, Tangen K. Dinoflagellates. In: Tomas CR. [ed.] Identifying Marine Diatoms and Dinoflagellates. New York; Academic Press: 1996. p. 387–598. [206] Steidinger KA. Implications of dinoflagellate life cycles on initiation of Gymnodinium breve red tides. Environ Lett 1975;9:129–39. [207] Steidinger KA, Truby EW, Dawes CJ. Ultrastructure of the red tide dinoflagellate Gymnodinium breve. I. General description. J Phycol 1978;14:72–9. [208] Faust MA, Gulledge RA. Identifying Harmful Marine Dinoflagellates. Smithsonian Institution, Contributions from the United States National Herbarium 2002;42:1–144. [209] Vargo GA. A brief summary of the physiology and ecology of Karenia brevis Davis (G. Hansen and Moestrup comb. nov.) red tides on the West Florida Shelf and of hypotheses posed for their initiation, growth, maintenance, and termination. Harmful Algae Volume 2009;8: 573–84. [210] Fu FX, Zhang Y, Warner ME, Feng Y, Sun J, Hutchins DA. A comparison of future increased CO2 and temperature effects on sympatric Heterosigma akashiwo and Prorocentrum minimum. Harmful Algae 2008;7:76–90. [211] Sun J, Hutchins DA, Feng Y, Seubert EL, Caron D A, Fu FX. Effects of changing pCO2 and phosphate availability on domoic acid production and physiology of the marine harmful bloom diatom Pseudo-nitzschia multiseries. Limnol Oceanogr 2011;56:829–40. [212] Matheson JR. The Effects of Ocean Acidification and Eutrophication on the Growth, Lipid Composition and Toxicity of the Marine Raphidophyte Heterosigma Akashiwo. University of Western Ontario – Electronic Thesis and Dissertation Repository 2014, Paper: 1983. [213] Pistocchi R, Guerrini F, Pezzolesi L, Riccardi M, Vanucci S, Ciminiello P, Dell’Aversano C, Forino M, Fattorusso E, Tartaglione L, Milandri A, Pompei M, Cangini M, Pigozzi S, Riccardi E. Toxin Levels and Profiles in Microalgae from the North-Western Adriatic Sea—15 Years of Studies on Cultured Species. Mar Drugs 2012;10:140–62. [214] Koray T. The occurrence of red tides and causative organisms in Izmir Bay. Ege Universitesi Fen Fakültesi Dergisi, Seri B 1984;1:75–83. [215] Nihayet Bizsel, Can Bizsel. New records of toxic algae Heterosigma cf. akashiwo and Gymnodinium cf. mikimotoi in the hypereutrophic Izmir Bay (Aegean Sea): Coupling between organisms and water quality parameters, Israel Journal of Plant Sciences 2002;50:33–44. [216] Labib W. Red tide occurrence in Alexandria (Egypt). A review. GFCM Workshop on Algal and Jellyfish Blooms in the Mediterranean and Black Sea. Istanbul, Turkey: 6th /8th October 2010. [217] Durgham H, Ikhtiyar S. First records of alien toxic algae Heterosigma akashiwo (Raphidophyceae) from the Mediterranean Coast of Syria. Arab Gulf Journal of Scientific Research 2012;30(1):58–60. [218] U.S. FDA (United States Food and Drug Administration). Fish and Fisheries Products Hazards and Controls Guidance, 3rd edition. Appendix 5 – FDA & EPA Safety Levels in Regulations and Guidance. June 2001. Available from http://www.fda.gov/downloads/Food/ GuidanceRegulation/UCM252448.pdf (accessed on 15/1/2015). [219] NZFSA (New Zealand Food Safety Authority). Animal products (specification for Bivalve Molluscan Shellfish, Notice 2006. Available from: http://www.foodsafety.govt.nz/elibrary/ industry/Animal_Products-Applies_Anyone.pdf (accessed on 15/1/2015). [220] FSANZ (Food Standards Australia New Zealand). Food Standard Code, Incorporating amendments up to and including Amendment 116, Standard 4.1.1, Primary Production and Processing Standards, Preliminary provisions, Standard 1.4.1, Contaminants and Natural toxicants, Issue 111. Available from http://www.comlaw.gov.au/Details/F2012C00285/Download (accessed on 15/1/2015).

Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

9 Gambierdiscus, the cause of ciguatera fish poisoning: an increased human health threat influenced by climate change 9.1 The genus Gambierdiscus Recent advances in population and species genetics for phytoplankton have revealed immense biodiversity at different taxonomic levels [1]. Vast numbers of species remain to be documented, aided by rapidly developing molecular methods [2]. To date, there are only approximately 160 described benthic (sand dwelling and epiphytic) dinoflagellates [3]. The first report by Yasumoto et al. [4] of the involvement of a benthic dinoflagellate in ciguatera fish poisoning (CFP) brought increased attention to this group. This species was described as Gambierdiscus based on the type species Gambierdiscus toxicus, from samples collected in the Gambier Islands, French Polynesia [5]. Species of the genus Gambierdiscus have now been recognized as the main producers of ciguatoxins (CTXs) and maitotoxins (MTXs) [6–12]. CFP is the most common nonbacterial illnesses associated with fish consumption [13], affecting between 50 000 and 500 000 people per year [14]. The ingestion of herbivorous and carnivorous fish that have orally accumulated effective levels of CTXs, and possible MTXs, can cause CFP in humans [15–17]. Recent reviews have illustrated the global increase in the frequency and intensity of harmful algal events [18, 19]. Despite being significantly underreported, CFP occurrence worldwide is increasing, with reports of a 60 % increase in CFP in the Pacific Islands over the last decade [20]. Once considered a monotypic taxon, new species of Gambieriscus are being discovered every year with evidence showing that each species might have its own characteristic toxin profile [9, 11, 12]. As in the case of other dinoflagellate genera such as Alexandrium or Karenia, the production or not of certain toxin groups appears to generally vary at the species level, rather than being consistent within the genus. For this reason, species of harmful algal bloom (HAB)-forming taxa are monitored, acting as early warning systems for shellfish and seafood safety. This review highlights the significant advances in the study of Gambierdiscus. We provide a summary of the morphology and phylogenetics of species of Gambierdiscus, their toxicology, distribution, chemistry and methods for the detection of CTXs and MTXs in seafood. The review further outlines the major gaps in our current understanding of Gambierdiscus and outlines goals for future research in this field.

274 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

9.2 Morphology and phylogenetics When originally described [5], Gambierdiscus was considered as a monotypic taxon, however, variability in the morphology, differences in ribosomal RNA (rRNA) genes, toxicity and physiological characteristics [5, 6, 21–27] led to the description of new species. Currently, 11 species of Gambierdiscus have been described, based on their distinct morphological and molecular genetic characteristics (Tab. 9.1). The following is an overview of the main morphological characteristics for each described species of Gambierdiscus. The original species descriptions consist of a comprehensive account of their characteristics. Gambierdiscus cells are large (60–100 μm), armoured, have a distinct plate pattern and fishhook shaped apical pore. Species are either anterio-posteriorly compressed (lenticular) or slightly laterally compressed (globular) (Fig. 9.1). The two globular species (G. yasumotoi and G. ruetzleri) can be distinguished from each other by cell size, size and shape of the 2󸀠 apical and 2󸀠󸀠󸀠󸀠 antapical plate and depth to width ratio, described in detail in Litaker et al. (2009) [28]. The remaining nine species are anterio-posteriorly compressed and broadly classified by either a narrow (G. australes, G. belizeanus, G. pacificus and G. excenreicus) or broad (G. polynesiensis, G. carolinianus, G. toxicus, G. caribaeus and G. carpenteri) 1p posterior intercalary plate. Among the species with a narrow 1p posterior intercalary plate, further distinguishing characteristics are either heavily areolated cell surface (G. belizeanus) or smooth cell surface species (G. australes, G. pacificus and G. excentricus). Species with a smooth cell surface can be distinguished by either having a hatchet-shaped 2󸀠 apical plate (G. pacificus) or more conventional rectangular shaped 2󸀠 apical plate (G. australes and G. excentricus). G. excentricus is at least 1.5 times wider and deeper than G. australes; further specifics distinguishing the two are described in detail in the original descriptions of the species [8, 11]. Species that have a broad 1p posterior intercalary plate can be further differentiated as having a rectangular shaped 2󸀠 apical plate (G. caribeaus and G. carpenteri) or a hatchet shaped 2󸀠 apical plate (G. toxicus, G. polynesiensis and G. carolinianus). G. toxicus can be further discerned by a pointed dorsal end to the 1p posterior intercalary plate. Further differences between G. polynesiensis & G. carolinianus are detailed in Litaker et al. [28]. G. caribeaus and G. carpenteri, both possessing a rectangular shaped 2󸀠 apical plate, are distinguished by the shape of 4󸀠󸀠 precingular plate, which is symmetric in G. caribaeus and asymmetric in G. carpenteri. The size and shape of the sulcal plates and various other specific morphological characteristics have also been described in the original descriptions of the species [7, 8, 11, 28, 29]. These features are straightforward to observe using light and scanning electron microscopy; however, within some species, a considerable amount of variability in features such as the size and shape of individual plates may be present. Another technique to identify different species of Gambierdiscus is to compare sequences that are known to be characteristic at the species level, such as regions of rRNA genes. Based on phylogenetic analysis of regions of the SSU (small ribosomal

Smaller size cell width less than 42 μm

(45.5 ± 3.3) × (37.5 ± 3.0) × (51.6 ± 4.9)

G. ruetzleri (Vandersea, Litaker, Faust, Kibler, Holland et Tester)

Narrow 1p plate heavily aerolated cell surface different 2󸀠 plate symmetry and size Narrow 1p plate smooth cell surface 2󸀠 hatch shaped Narrow 1p plate smooth cell surface 2󸀠 rectangular shaped cell size bigger than G. australes (1.5 times wider and deeper)

(61.7 ± 3.1) × (60.0 ± 4.5) × (48.1 ± 4.2) (58.5 ± 3.9) × (53.6 ± 4.1) × (40.4 ± 3.6) (97.8 ± 8) × (83 ± 10) × (37 ± 3)

G. belizeanus (Faust)

G. pacificus (Chinain et Faust)

G. excentricus (Fraga)

Anterio-posteriorly compressed species

Larger cell size cell width larger than 42 μm

Morphological characteristics (plate formula)

(56.8 ± 5.6) × (51.7 ± 5.6) × (62.4 ± 4.3)

Cell size (μm) (depth × width × length)

G. yasumotoi (Holmes)

Globular species

Species

Tab. 9.1: Taxonomic and genetic identifications of different species of Gambierdiscus.

D1-D3 LSU: HQ877874, JF303063, JF303065-71 D8-D10 LSU: JF303073-76

SSU: EF202861-65 D1-D3 LSU: EF202944-47 D8-D10 LSU: EF498012-13, EF498015-16

SSU: EF202876-77 D1-D3 LSU: EF202940-43 D8-D10 LSU: EF498028-34

SSU: EF202853-60 D1-D3 LSU: EF202962-64 D8-D10 LSU: EF498081-85

SSU: EF202846-52 D1-D3 LSU: EF202965-68 D8-D10 LSU: EF498087-89

Genetics

[11]

[8, 28]

[28, 29]

[28]

[7, 28]

References

9 Gambierdiscus, the cause of ciguatera fish poisoning |

275

Morphological characteristics (plate formula) Narrow 1p plate smooth cell surface 2󸀠 rectangular shaped smaller than G. excentricus Broad 1p plate 2󸀠 Rectangular shaped Symmetric 4󸀠󸀠 Broad 1p plate 2󸀠 Rectangular shaped Asymmetric 4󸀠󸀠 Bad 1p plate 2󸀠 Hatchet shaped Dorsal end 1p pointed Broad 1p plate 2󸀠 Hatchet shaped Dorsal end 1p oblique smaller cell size than G. carolinianus

Cell size (μm) (depth × width × length) (72.5 ± 3.8) × (63.4 ± 5.0) × (38.7 ± 3.8) (82.2 ± 7.6) × (81.9 ± 7.9) × (60 ± 6.2) (81.7 ± 6.4) × (74.8 ± 5.9) × (50.2 ± 6.1) (93 ± 5.5) × (83 ± 2.3) × (54 ± 1.5) (66.3 ± 3) × (60.5 ± 5.9) × (44.3 ± 5.1)

Species

G. australes (Faust et Chinain)

G. caribaeus (Vandersea, Litaker, Faust, Kibler, Holland et Tester)

G. carpenteri (Vandersea, Litaker, Faust, Kibler, Holland et Tester)

G. toxicus (Adachi et Fukuyo) Chinain, Faust, Holmes, Litaker et Tester)

G. polynesiensis (Chinain et Faust)

Tab. 9.1 (continued)

SSU: EF202902-07 D1-D3 LSU: EF202976-82 D8-D10 LSU: EF498076-80

SSU: EF202878-90 D1-D3 LSU: EF202951-61 D8-D10 LSU: EF498017-27

SSU: EF202908-13 D1-D3 LSU: EF202938-39, EF202984 D8-D10 LSU: EF498038-44

SSU: EF202914-28 D1-D3 LSU: EF202929-37, EF202983, EF202985 D8-D10 LSU: EF498045-71

SSU: EF202891-96 D1-D3 LSU: EF202969-72 D8-D10 LSU: EF498072-74

Genetics

[8, 28]

[5, 6, 25, 28]

[28]

[28]

[8, 28]

References

276 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

Broad 1p plate 2󸀠 Hatchet shaped Dorsal end 1p oblique larger cell size than G. polynesiensis

(78.2 ± 4.8) × (87.1 ± 7.1) × (51.4 ± 5.2)

G. carolinianus (Vandersea, Litaker, Faust, Kibler, Holland et Tester)

Not described Not described

Not described

Not described Not described

Gambierdiscus ribotype 1

Gambierdiscus ribotype 2

Gambierdiscus sp. type 1

Gambierdiscus sp. type 2

Gambierdiscus sp. type 3

Genetically described phylotypes

Morphological characteristics (plate formula)

Cell size (μm) (depth × width × length)

Species

Tab. 9.1 (continued)

SSU: AB764296-300 LSU D8-D10: AB765923-24

SSU: AB764277-96 LSU D8-D10: AB765913-18

SSU: AB64229-76, AB605799-800, AB605811-12 LSU D8-D10: AB765908-13

D8-D10 LSU: GU968499-500, GU968503, GU968505, GU968507-11

D8-D10 LSU: GU968512-20, GU968523

SSU: EF202897-EF202901 D1-D3 LSU: EF202973-75 D8-D10 LSU: EF498035-37

Genetics

[32]

[30, 32]

[30, 32]

[31]

[31]

[28]

References

9 Gambierdiscus, the cause of ciguatera fish poisoning |

277

278 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

Gambierdiscus ruetzleri

Gambierdiscus belizeanus

4” 5” 3’ 2’ 3” 6” Po 2” 1’

4”

6”

7” 1”

Gambierdiscus yasumotoi

1’

2’

3” 3’

5”

Gambierdiscus polynesiensis 4”

5” 3’ 2’ 3”

2’

3’

5”

1’

1’

2”

7”

1”

Gambierdiscus carolinianus 4”

3”

2’

Po

6” 7”

2” 1”

3’

5”

7”

2”

1’

1”

7”

4” 5”

3’ 6”

Gambierdiscus australes

1’

2”

Gambierdiscus excentricus

4”

2’ 3” Po

4” 2’

5” 3’

3” Po

6”

Gambierdiscus pacificus

3” Po

2” 6”

4”

6” Po 2” 1’ 1”

4”

2’ Po

5” 3’

Gambierdiscus caribaeus

3”

2’

3”

Po

6” 1’

5”

2”

3’ 6”

Po 1’

2”

7” 1” Gambierdiscus carpenteri

4” 2’

6”

1’ 7”

4”

3”

Po

5” 3’

Gambierdiscus toxicus

2”

2’ 3’

5”

3” Po

1”

1’

6” 7”

1”

Fig. 9.1: Comparative line drawings of the epitheca for 11 Gambierdiscus species. Sale bar equals 50 μm. Modified from Litaker et al., 2009 [28].

9 Gambierdiscus, the cause of ciguatera fish poisoning

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subunit) rDNA, LSU (large ribosomal subunit) rDNA and ITS (internal transcribed spacer) rDNA, the genus Gambierdiscus is monophyletic [8, 11, 25, 28, 30–32]. Further, the lenticular and globular species form two distinct clades. Phylogenetic analysis has shown that the two globular species (G. ruetzleri and G. yasumotoi) diverged relatively early in the evolution of the genus Gambierdiscus [28, 32]. Also, G. ruetzleri and G. yasumotoi are the two most closely related species in the genus. Based on LSU rDNA D08-D10 sequences, the mean p distance within species is 0.002±0.002, and between species is 0.121 ± 0.036 (calculated based on sequences from 10 species/phylotypes) where minimum p distance between G. ruetzleri and G. yasumotoi is 0.007 [32]. Using SSU rDNA sequences, the mean p distance within species is 0.003 ± 0.002 and between species is 0.139 ± 0.042 (calculated based on sequences from 10 species/ phylotypes) where minimum p distance between G. ruetzleri and G. yasumotoi is 0.004 [32]. These statistics are indicative of putative unknown species and can be very useful in cases where morphological, physiological or other data is not yet available, or a strain is not present in culture. Based on D8-D-10 LSU rDNA phylogenetic analysis, two new putative phylotypes Gambierdiscus ribotype 1 and Gambieridsuc ribotype 2 were reported [31], as the two clusters/clades separated from the others and their genetic distances equalled or exceeded those among the 11 described species [31] (Tab. 9.1). Similarly, three new putative species/phylotypes of Gambieridiscus (Gambieridsucs sp. type 1, type 2 and type 3) have been described from Japan based on differences in the regions D8-D10 of the LSU and SSU rDNA [30, 32] (Tab. 9.1). In this case, the p distances between these two novel clades and known species of Gambierdiscus were larger than those separating G. yasumotoi from G. ruetzleri. Although the genetic data indicates that these phylotypes may be new species, their morphological circumscriptions are needed to support their status as new species. As sampling around the world becomes more intensive, it is likely that new species of Gambierdiscus will be described.

9.3 Geographic distribution and abundance Gambierdiscus is widely distributed in coastal zones at tropical and subtropical latitudes. However, the distribution of species of Gambierdiscus is still poorly understood, as the discrimination of different species of Gambierdiscus has only occurred recently (Tab. 9.2).

N/K

Yes

N/K

N/D [10]

HELA-positive [12]

N/K

HELA-positive, MBA-positive [8, 10, 32] MBA-positive [8]

MBA-positive [8] HELA-positive [12]

RBA-positive [9]

MBA-positive [7]

MBA-positive [8, 10, 32], RBA-positive [9] MBA-positive [8]

MBA-positive [8], RBA-positive [9] N/K

Belize [29], Florida [28] , Mexican Caribbean [50], Malaysia [138], Pakistan [139], Queensland, Australia (murray unpubl. Data), St. Barthelemy Island – Caribbean [31]

Singapore [7], Japan [32], Mexican Caribbean [50], Queensland, Australia (murray unpubl. Data), Nha Trang – Vietnam [137]

French Polynesia [8], Japan [32], Cook Islands [10], Hawaii USA [28], Pakistan [139]

French Polynesia [8], Marshall Islands & Society Islands Micronesia [31], Kota Kinabalu and Sipandan Island [140], Nha Trang – Vietnam [137]

French Polynesia [8], Canary Islands [11], Pakistan [139], Nha Trang – Vietnam [137]

Florida, Belize – Caribbean, Tahiti, Palau, Hawaii [28], Flower Gardens – Gulf of Mexico, Osho Rios – Jamaica [12], Bahamas, Grand Caymam Island, Tol-truk Micronesia [31], Jeju Island Korea [141]

G. yasumotoi

G. australes

G. pacificus

G. polynesiensis

G. caribaeus

N/K

Yes [9]

N/K

N/K

N/K

N/K

N/K

N/K

N/K

G. belizeanus

N/K

MBA-positive [8]

MBA-negative [8], RBA-positive [9]

Tahiti, French Polynesia [5, 8], Mexican Caribbean [50], New Caledonia, Reunion Island, Indian Ocean [8], Nha Trang – Vietnam [136, 137]

G. toxicus

CTX

MTX

CTX

MTX

LC-MS

Toxicity various assays

Geographical distribution

Species

Tab. 9.2: Geographic distribution and toxicity of different Gambierdiscus species.

280 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

N/K N/K MBA-positive [32] MBA-negative [32] MBA-negative [32]

Belize – Caribbean [31]

Belize – Caribbean, Martinique – Caribbean [31], Puerto Rico [12]

Japan [32]

Japan [32]

Japan [32]

Gambierdiscus ribotype 1

Gambierdiscus ribotype 2

Gambierdiscus sp. type 1

Gambierdiscus sp. type 2

Gambierdiscus sp. type 3

N/K

N/K

N/K

N/K

N/K

N/K

N/K

HELA, human erythrocyte lysis assay,

MBA-positive [32]

MBA-negative [32]

MBA-positive [32]

HELA-positive [12]

N/K

NCBA-positive [11]

HELA-positive [12]

N/K

N/K

N/K

N/K

N/K

N/K

N/K

N/K

N/K

N/K

MTX

NCBA, neuro-2a

NCBA-positive [11]

Canary Islands [11]

G. excentricus

RBA, receptor-binding assay,

N/K

North Carolina, USA, Belize – Caribbean [28]

G. ruetzleri

The abbreviations are: N/K, not known, N/D, not detected, MBA, mouse bioassay, cell binding assay.

N/K

Belize, Guam, Fiji [28], Hawaii [31], Dry Tortugas – Florida, Flower Gardens – Gulf of Mexico, Osho Rios – Jamaica [12]

G. carpenteri

HELA-positive [12]

HELA-positive [12]

N/K

North Carolina, USA, Atlantic ocean [28], Bermuda, Mexico [31], Puerto Rico, Flower Gardens – Gulf Of Mexico, Osho Rios – Jamaica, Crete – Greece [12]

N/K

CTX

G. carolinianus

LC-MS MTX

CTX

Toxicity various assays

Geographical distribution

Species

Tab. 9.2 (continued)

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9.3.1 The Pacific and Indian Ocean Regions Gambierdiscus is named after the Gambier Islands in French Polynesia, where it was first identified, [5], and since then, G. toxicus, G. belizeanus, G. yasumotoi, G. australes, G. pacificus, G. polynesiensis, G. caribaeus and G. carpenteri have been reported from various Pacific islands, Hawaii, Australia, Southeast Asia and the Northern Indian Ocean (Tab. 9.2). Recently, three genetically distinct species from coastal and temperate waters of Japan were reported [32] (Tab. 9.2). In addition, Gambierdiscus has been reported from the Philippines [16], Hong Kong [33], Indonesia [34] and Mauritius [35], although species diversity in these areas is not known. Gambierdiscus has also been reported from the Mexican Pacific coast [36] and regions around Madagascar [37], where cases of CFP have also been previously reported [38, 39].

9.3.2 The Atlantic Ocean Region Early accounts of Gambierdiscus “look-alike” species date back to 1948 from Cape Verde Islands [40] and 1979 from Key Largo, Florida [41]. So far, G. toxicus, G. belizeanus, G. yasumotoi, G. polynesiensis, G. caribaeus, G. carolinianus, G. carpenteri, G. ruetzleri, G. excentricus, Gambierdiscus ribotype 1 and Gambierdiscus ribotype 2 have been reported from the east coast of the USA, Caribbean and the Mediterranean regions (Tab. 9.2). There are many other regions where Gambierdiscus has been reported; however, the exact species are yet to be determined. These include Cyprus, Rhodes, Saronikos Gulf [42, 43], French West Indies [44], Cuba [45] and Veracruz [46]. Other confirmed reports of Gambierdiscus occurrence in Central and South America in the literature are from Costa Rica and Brazil (M. Montero pers. comm. in [47]). From Africa, there has been only one direct observation of Gambierdiscus, from the coast of Angola [48]; however, CFP cases from the west coast (Canary Islands and Cameroon) [49] of Africa have been reported, indicating the presence of Gambierdiscus in that region. Certain species of Gambierdiscus have been designated as being endemic to either the Pacific or the Atlantic regions [31, 48]. So far, G. australes and G. pacificus have only been reported from the Pacific, and G. ruetzleri, G. excentricus and Gambierdisucs ribotype 1 and 2 are only reported from the Atlantic region (Tab. 9.2). G. belizeanus, G. caribeaus, G. carpenteri and G. carolinianus are widely distributed in the Atlantic and Pacific Oceans [28, 31, 48]. G. yasumotoi is widely distributed in the Pacific; however, there is only one report of its occurrence in the Mexican-Caribbean [50], which was reported before the discovery of the other globular species G. ruetzleri, which is widely distributed in the Atlantic region [28]. The distribution of G. toxicus needs to be refined, due to numerous misidentifications in the literature. G. polynesiensis is widespread in the Pacific (Tab. 9.2) with only one confirmed report from the Canary Islands in the Atlantic region [11]. Both, Litaker et al, 2010 and Berdalet et al.,

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2012 [31, 48] mention that none of the Pacific-specific species have been observed in hundreds of field samples analysed from Atlantic regions (Caribbean/Gulf of Mexico/West indies/Southeast US coast from Florida to North Carolina). The absence of Atlantic-specific species in the Pacific region has not been confirmed, as the majority of the vast Pacific region remains unexplored. As under-sampling and underreporting have occurred worldwide, but particularly in the Pacific region, much more work needs to be undertaken in order to determine whether endemism or restricted distributions exist in species of Gambierdiscus. While multiple species of Gambierdiscus can co-occur in one region, equally, there are regions from where only one species has been reported. For example, in Heron Island (Queensland, Australia) there are at least three species of Gambierdiscus that co-occur, however further south in Merimbula, New South Wales only G. carpenteri is known to occur (Murray, unpublished data). Localized benthic blooms of Gambierdiscus have been noted in the literature from both the Pacific and Atlantic regions [51–54]. Cell densities in such blooms can range from anywhere between 10 000 to 100 000 cells g−1 wet weight algae [31]. There are no accurate estimates of cell densities at which a Gambierdiscus bloom leads to a CFP epidemic. The onset of CFP may also depend on other factors, such as the fact that different species of Gambierdiscus have varying toxicities. For example, in 2010, an unidentified species of Gambierdiscus was reported in Greece, however no CFP outbreaks have been reported there [55]. In most habitats where species of Gambierdiscus occur, cell densities are below 1000 cells g−1 wet weight algae [31], however in some environments Gambierdiscus spp are known to occur year-round at such cell densities [51]. A constant exposure of low densities of cells could also lead to a build up of CFP-related toxins in fish. Much more research needs to be done in order to understand the relationship between Gambierdiscus abundance and CFP outbreaks. This is particularly challenging, as benthic dinoflagellates inhabit areas where quantitative sampling of microbial eukaryotes is not straightforward, for example, in sediments and on the surface of dead corals. Also, Gambierdiscus cell distribution can be very patchy, even over small distances, making it hard to estimate mean Gambierdiscus cell densities over a larger area [31, 44, 56].

9.4 CTXs and MTXs CTXs are sodium channel activators, particularly affecting the voltage sensitive channels located along the nodes of Ranvier (peripheral nerve cells) [57–59]. When the sodium channels are activated, there is a massive influx of Na+ ions, resulting in cell depolarization [57–59]. This leads to the onset of spontaneous action potentials in effected cells, causing various symptoms in humans. Symptoms can include but are not limited to gastrointestinal, neurological and in cases of severe intoxication sometimes cardiovascular [17], and can vary depending on geographical region [59, 60]. This can be due to the structural differences of CTXs in different regions, therefore it

284 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

is very important to characterize CTXs from Pacific, Caribbean and the Indian Oceans. Local understanding of CTX accumulation patterns in different fish species can also help prevent CFP. However, the accurate identification of exact congeners of CTXs is necessary, in order to understand the toxicology and evaluate the local risks of CFP. Structurally, CTXs are thermostable, cyclic polyether ladders, which are liposoluble (Fig. 9.2). They have been isolated from fish and different species of Gambierdiscus (Tab. 9.3). Based on their origin and differences in the structure of these toxins, they are divided into P-CTXs (Pacific Ocean), C-CTXs (Caribbean region) and I-CTXs (Indian Ocean). Due to their structural differences, P-CTXs are further divided into type I and type II, as suggested by Legrand et al. [61]. Type I P-CTXs have 13 rings and 60 carbon atoms [62–65]. This category consists of the first CTX to be fully structurally described as CTX1B [62] (or CTX-1 as described by Lewis et al. 1991, [63]) from moray eels, which is the principal toxin in the carnivorous fish from the Pacific [62, 63]. Two other type I P-CTXs, i.e. CTX-2 and CTX-3, were also described from the same extracts; they have slight variations in their structures leading to different toxicities in mice [63] (Tab. 9.3). It has also been suggested that CTX-1, CTX-2 and CTX-3 may be derived from dinoflagellate precursors known as CTX-4A and CTX4B (also named as GTX-4B in [62]) [64, 65]. Recently, CTX-4A and CTX-4B have been isolated from G. polynesiensis culture extracts [9]. CTX3C is a type II P-CTX with 13 rings, 57 carbon atoms and was first isolated from cultures of Gambierdiscus sp. [66] and later from G. polynesiensis [9]. Two more congeners of CTX3C called as 49-epi-CTX-3C (also called as CTX-3B in [9]) and M-seco-CTX-3C have also been isolated from Gambierdiscus sp. [66] and G. polynesiensis [9]. Later, 2 new type II P-CTXs, i.e. 2,3 dihydroxyCTX3C (also called as CTX2-A1) and 51-hydroxyCTX3C, were isolated from Moray eel [67] that might be oxygenated metabolites of CTX3C [65]. Caribbean CTXs are slightly bigger than P-CTXs and have 14 rings and 62 carbon atoms [68–71]. Many congeners of C-CTXs have been isolated from carnivorous fish, including C-CTX1, C-CTX-2, C-CTX-1141, C-CTX-1127, C-CTX-1143, C-CTX-1157, C-CTX-1159 [68–71]. Unlike P-CTXs, there have been no reports of C-CTXs originating from Gambierdiscus sp. However, recently G. excentricus has been identified as a major CTX producer in the Caribbean [11], and CTXs from this strain are being characterized. Recently 4 CTXs (I-CTX-1, I-CTX-2, I-CTX-3, I-CTX-4) have been isolated from carnivorous fish from Indian Ocean and have higher molecular ion masses than P-CTXs and C-CTXs [55, 72, 73]. However, their structures need to be elucidated [72, 73]. I-CTX-1 is toxic to mice via intraperitoneal injection [73]. Based on mouse bioassays (MBA), different congeners of CTXs can have variable toxicities (Tab. 9.3), however this needs to be further validated as well.

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Tab. 9.3: Different congeners of CTXs and MTXs. Origin

Toxin Name

Molecular Ion [M + H]+

Source

Toxicity*

CTX1B [62], CTX-1[63]

1111.6 [62, 63]

Moray eel (Gymnothorax javanicus) [62] Moray eel (Lycodontis javanicus, Muraenidae) [63]

CTX1B: 0.35 μg/kg [62] CTX-1: 0.25 μg/kg [63]

CTX-2

1095.5 [63]

Moray eel (Lycodontis javanicus, Muraenidae) [63]

2.3 μg/kg [63]

CTX-3

1095.5 [63]

Moray eel (Lycodontis javanicus, Muraenidae) [63]

0.9 μg/kg [63]

CTX4A

1061.6 [65]

Gambierdiscus sp. [65] G. polynesiensis [9]

12 μg/kg [9]

CTX4B

1061.6 [65]

Gambierdiscus sp. [65] G. polynesiensis [9]

20 μg/kg [9]

CTX3C

1023.6 [66]

Gambierdiscus sp. [66] G. polynesiensis [9]

2.5 μg/kg [9]

49-epi-CTX-3C

1023.6 [9]

Gambierdiscus sp. [66] G. polynesiensis [9]

8 μg/kg [9]

M-seco-CTX-3C

1041.6 [9]

Gambierdiscus sp. [66] G. polynesiensis [9]

10 μg/kg [9]

C-CTX-1

1141.6 [68, 70]

Horse-eye jack (Caranx latus)

3.6 μg/kg [68]

C-CTX-2

1141.6 [68, 70]

Horse-eye jack (Caranx latus)

Toxic [68]

I-CTX-1

1141.6 [73]

Red bass (Lutjanus bohar) and red emperor (Lutjanus sebae) [73]

Toxic [73]

MTX-1

3422 [74, 82]

Gambierdiscus sp. [82]

0.05 μg/kg [74]

MTX-2

3298 [82]

Gambierdiscus sp. [82]

0.08 μg/kg [82]

MTX-3

1060 [82]

Gambierdiscus sp. [82]

Toxic [82]

Ciguatoxins Pacific (type I)

Pacific (type II)

Caribbean

Indian

Maitotoxins Pacific

* LD50 doses calculated via i. p. injection in mice.

286 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

CH3

OH O

CH3

O

O

R2

O

HO

O

O

CH3

O O

O

O

OH O H3C

Type I CTX backbone Ciguatoxin, R1;R2 P-CTX-1, OH; CH(OH)CH2OH P-CTX-2, H; CH(OH)CH2OH P-CTX-4B, H; CH=CH2

CH3

O

O HO

H3C

CH3

O O

O

O

R1

O

O

O

OH O

Type II P-CTX-3C

CH3

O HO

CH3

O O

O

O

OH CH3 O

O Type III C-CTX-1 H

OH

O

H

H O

OH

H

H O

H

O

O

H

O

O

OH

OH

O

H

H

O

O H

O

H3C H3C

H

O HO

R1

O

O

O

O

H3C

CH3

OH O

O

H3C

HO O

O

CH3

OH O

O

O

H

H

H

H

H

O

O H O H O OH HO

HO O

H O OH

OSO3Na

OH H HO

O H H OH

H

O O

H

OH

NaO3SO OH OH

H

H

H OH

O H

H OH

OH

H

O

H

O H

H

OH

H H

OH

H OH

O

HO H O

O H

H O

OH H

O H

O

H

H HO OH

H

O

H

OH

HO H

O

OH

OH

H

Maitotoxin-1 Fig. 9.2: Structure of Ciguatoxins (CTX) and Maitotoxin-1. P-CTX-1, P-CTX-2 and C-CTX-1 were derived from fish and P-CTX-3C, P-CTX-4B and Maitotoxin-1 were derived from Gambierdisucs spp.

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Maitotoxins (MTXs) are one of the largest non-proteinous and highly toxic natural products known [74, 75]. This polyether ladder type compound was first discovered as a water-soluble toxin in the guts of herbivorous fish Acanthurids (surgeonfish) in 1976 [76]. In the 1990s, stereoscopic studies and partial synthesis were used to determine the structural elucidation and stereochemistry of the extraordinary complex and large MTX [74, 77–81]. Simultaneously, Holmes & Lewis described two large (MTX-1, MTX-2) and one small MTX (MTX-3) from different strains of Gambierdiscus sp. isolated from Queensland, Australia [82] (Tab. 9.3). MTX-1 from this study may have been the MTX originally described from guts of Acanthurids, however it is not clearly proven. When compared to other natural toxins, MTX is a highly potent calcium channel inhibitor (LD50 0.05 μg/kg, i.p., mice), only exceeded by a handful bacterial proteinous toxins [74, 75]. Despite its high level of potency, the complete mode of action and the primary target of MTX in mammalian cells have not yet been fully elucidated. In fact, the activation of voltage dependent calcium channels induced via MTXs is a secondary effect of membrane depolarization (for review and more details see [83]). Recently, it has been reported that the biophysical mechanisms of pacific MTXs are different to Caribbean MTXs [84]. Whether this is due to a structural difference is not known, as the Caribbean MTXs have not been fully characterized. Although MTX appears to have a low tendency of accumulating in fish flesh, as compared to stomach or intestines [76], its possible role in CFP cannot be disregarded, as eating non-eviscerated fish is a common practice in many Pacific Island nations. The sulphate esters in the structures of MTXs make it amenable to detect and quantify MTX by LC_ESI_MS (Liquid chromatography-electronspray ionisation-Mass spectrometry) (T. Harwood, pers. comm.), and Solvolysis (desulphonation) reduces the toxicity of MTXs significantly, at least 100-fold [85]. However, more research is essential to understand the exact role of MTXs in CFP including its mode of action and target in mammalian cells. Cyclic polyether ladders are almost exclusively known to be produced by dinoflagellates. Other than CTXs and MTXs, this class of secondary metabolites also includes Brevetoxins (BTXs), produced by Karenia spp [86] and Yessotoxins (YTXs), produced by a wide array of dinoflagellates including Lingulodinium polydrum [87], Gonyaulax spinifera [88] and Protoceratium reticulatum [89]. Based on their high structural similarities, the synthesis of these compounds likely involves common biosynthetic mechanisms [90–92]. Stable-isotope labeling of precursors to elucidate the biosynthesis pathway of CTXs and MTXs has never been performed. However, precursor studies to reveal the biosynthesis pathways of BTXs and YTXs have indicated the polyketide origin of these cyclic polyether ladders [93, 94]. Several schematic pathways involving different enzymes have been suggested and are detailed in Kalaitzis et al. [95] and Kellmann et al. [94]. It is speculated that the biosynthesis involves the normal polyketide synthase (PKS) enzyme complex with a few additional enzymes, i.e. expoxidases and thioesterases [96]. Essential domains present in the PKS are: acyltransferase domain (AT); β-ketosynthase domain (KS); and acyl carrier protein (ACP) [97]. In addition, PKS can include β-ketoacyl reductase (KR), enoyl reductase (ER) and dehydrogenase (DH)

288 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

domains [97]. In the past 10 years, a few genes that encode the essential domains of the PKSs, particularly KS domains in dinoflagellates, have been identified for the first time. However, with the availability of next generation sequencing tools, a few candidate genes encoding KS and KR domains in Karenia brevis have been associated with biosynthesis of BTXs [98]. A recent study published a comprehensive transcriptome library of Lingulodinium polydrum for which genes encoding KS domains were reported, however no link between these genes and YTX production has been established [99]. In the past, a few studies have identified genes encoding KS domains in Amphidinium sp. [100], which produces numerous macrolides (cyclised linear polyethers) such as amphidinolides. No studies have been done to identify genes involved in CTX and MTX biosynthesis. However, an extensive marine microbial eukaryote transcriptome project, undertaken by the Moore Foundation, is in the process of sequencing 652 trancriptomes, which includes 2 strains of Gambierdiscus species. Analysis of data obtained for such diverse arrays of dinoflagellate species may shed light on the genes involved in secondary metabolite synthesis in dinoflagellates.

9.5 Toxicity of different species of Gambierdiscus There is clear evidence that Gambierdiscus species produce CTXs and/or MTXs [24, 62, 66, 82, 101]. However, many wild and cultured strains of Gambierdiscus have not been found to produce detectable amounts of CTXs [23, 102]. Unfortunately, most of these studies describe the identity of the cultures as Gambierdiscus toxicus, since it was the only known species of Gambieriscus at that time. It is imperative to study the toxin profile of all of the species and genotypes now known (Tab. 9.2). Tab. 9.2 provides the data available on the toxicity of each species of Gambierdiscus detected via various assays and LC-MS (liquid chromatography-mass spectrometry)-based detection. In 2010, Chinain et al. [9] described the toxin profile of G. polynesiensis based on LC-MS analysis and receptor binding assay (RBA). This species produces both Type 1 (CTX-4A, CTX-4B) and Type 2 P-CTXs (CTX-3C, M-seco-CTX-3C, 49-epiCTX-3C), however P-CTX-3C was the major toxin produced by this species. Two different strains of G. polynesiensis were tested and found to produce same suite of toxins, in different proportions [9]. Similar results were found in a study of 56 strains of six different species (G. belizeanus, G. caribaeus, G. carolinianus, G. carpenteri, Gambierdiscus ribotype 2) over a period of two years, using the human erythrocyte lysis assay (HELA) [12]. The intraspecific toxicity varied slightly among different strains of same species, however the level of toxicity of each strain remained unchanged over the period of the study [12]. HELA assay toxicity is indicative of MTX production by the species. The water-soluble fraction of the extracts of G. polynesiensis has been found to be toxic via MBA [8], indicating the presence of MTXs. However, the toxins that produced this effect have not been characterized from this strain. Another species from the Caribbean, G. excentricus, may produce CTXs and MTXs (as determined via Neuro-2a cell based assay) [11], how-

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ever the exact toxin profile needs to be verified via LC-MS analysis. The toxicities of the liposoluble and water-soluble fractions of G. australes extracts, isolated from the Cook Islands, were found to be toxic via MBA, indicating the presence of CTXs and MTXs [10]. However no CTXs were detected via LC-MS analysis [10]. Another strain of G. australes from French Polynesia tested positive for CTXs via the RBA; however, the level of toxicity was low when compared to G. polynesiensis [9]. These results are intriguing and require further analysis. While bioassays are important to determine toxicity, only LC-MS-based analysis techniques can determine the exact toxin profile of different species of Gambierdiscus. As we only know the partial toxin profiles of two species of Gambierdiscus via LC-MS-based techniques, this area of research needs urgent attention.

9.6 Detection of CTXs and MTXs in seafood Originally, CFP was derived from the word “cigua”, used by native Cubans to describe a turban-shelled snail and implicated in an outbreak of the sickness in Spanish explorers to Cuba in the 1500s [103]. The occurrence of CTX in the turban snail Turbo argyrostoma has been confirmed [104]. However, to date, the majority of cases reporting occurrences of CFP have followed consumption of large reef fish (e.g. [105–108]). This circumstance has been a critical factor in the diagnosis of the disease, as in many cases there has been no fish sample retained for chemical verification or the appropriate test facilities have not been available. Although hundreds of cases of CFP have been documented worldwide, it is estimated that less than 20 % of actual cases have been reported [109]. There is a high likelihood of misdiagnosis for CFP. The number of documented symptoms, which are in excess of 175 [17], may vary depending on portion size [110], individual susceptibility or accumulation of toxin with age [15, 111] and could also be associated with other illnesses (e.g. decompression sickness [112], chronic fatigue syndrome, multiple sclerosis [113, 114] and brain tumors [113]). The number of fish species implicated in ciguatera outbreaks is suggested to be of the order of several hundred (Halstad, 1978, FAO, 2004). However, with the above limitations and the absence of a reliable, commercially available test kit, it is difficult to express an exact figure. While carnivorous fish are the main culprits, herbivorous fish (e.g. surgeonfish and parrotfish), a key component of the toxic food chain [115, 116], have also been linked to CFP outbreaks. Tab. 9.4 provides a summary of over 90 fish species and other marine fauna that have tested positive for CTXs, from ciguatera prone regions and following reported outbreaks.

C-CTX-1 [71, 107]

CTX – positive [150]

The Bahamas [147], West Africa [49], Florida Keys, USA [107], French West Indies [71], St. Barthelemy, Caribbean Sea [144, 148], Guadeloupe [148], French Polynesia [149]

Hervey Bay, Queensland, Australia [150]

Sphyraena barracuda (Great barracuda)

Sphyraena jello (Pickhandle barracuda)

TLC & MBA [150]

Cat BA [149], Chick BA [148], MQBA [149], MBA [149], N2A [147]

LCMS/MS [146], BSBA [126], MGBA [126], ELISA [129, 130], N2A [129, 130, 146]

C-CTX-1 [146]

Canary Islands [146], Hawaii [129, 130], St. Thomas, Carribean Sea [126]

Seriola rivoliana (Almaco jack-Kahala)

BarracudaC

LCMS/MS [145], UPLC/MS) [143]

C-CTX-1 [145], CTX-1B [143], CTX-3C and CTX analogues from Carribean or Indic waters) [143]

Selvagens Islands (Madeira Arquipelago) [143], West Africa [145]

Seriola fasciata (Lesser amberjack)

UPLC/MS [143], HPLC/MS [134], TLC [144], BSBA [126], MGBA [124, 126], MBA [128, 131, 144], S-EIA [131], SPIA [143], RIA [128], ELISA [128, 129], N2A [129, 142], RBA [134]

Method of detection

C-CTX-1 [134], CTX-1B [143], CTX-3C and CTX analogues from Carribean or Indic waters [143]

CTX (if detected)

Canary Islands [142], Selvagens Islands (Madeira Arquipelago) [143], Hawaii [124, 128, 129, 131], Haiti [134], St. Barthelemy, Caribbean Sea [144], St. Thomas, Carribean Sea [126]

Source

Seriola dumerili (Greater amberjack- Kahala)

AmberjackC

Latin name (Common name)

Tab. 9.4: Different congeners of Ciguatoxins detected by various assays in seafood and other animals.

290 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

California [132]

South Taiwan [151]

Sphyraena sp. (Barracuda)

Sphyraena spp. (Barracuda fish eggs)

Mulloidichthys auriflamma (Goldstriped goatfish)

CTX – positive [132]

CTX – positive [149]

French Polynesia [149]

Monotaxis grandoculis (Big eye bream)

Hawaii [132]

CTX – positive [149]

French Polynesia [149]

Lethrinus miniatus (Trumpet emperor)

GoatfishC

CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Lethrinus olivaceus (Longface emperor)

S-EIA [132], SPIA [132]

Cat BA [149], MQBA [149], MBA [149]

Cat BA [149], MQBA [149], MBA [149]

RBA [54]

HPLC/MS [67, 152], HPLC/HNMR [62, 63, 120], TLC [153], DLBA [125], MBA [67, 152, 153]

CTX-1 [62, 63], CTX-4B [62, 63], CTX-2 [63], CTX-3 [63], P-CTX-1 [152], P-CTX-2 [152], P-CTX-3 [152] and analogues of CTX 3C: 2,3-dihydroxyCTX3C and 51-hydroxyCTX3C [67]

Tuamotu Archipelago and Tahiti (French Polynesia) [62, 120, 125], Tarawa, Republic of Kiribati, central Pacific Ocean [152], Hawaii [153]

Gymnothorax javanicus (Moray eel)

Emperor breamC

TLC [144], MBA [144]

CTX – positive [144]

MBA [151], N2A [151]

S-EIA [132], SPIA [132], N2A [132]

Method of detection

St. Barthelemy, Caribbean Sea [144]

CTX – positive [151]

CTX – positive [132]

CTX (if detected)

Gymnothorax funebris (Green Moray)

EelC

Source

Latin name (Common name)

Tab. 9.4 (continued)

9 Gambierdiscus, the cause of ciguatera fish poisoning |

291

CTX – positive [155]

Rep. of Vanuatu [155]

Hippopus hippopus (Giant Clam)

CTX – positive[54, 129, 149] P-CTX-1 [60, 156, 157] CTX – positive [133] CTX – positive [158]

Nuku Hiva (Marquesas) [54], Hawaii [129], French Polynesia [149]

Fiji [60, 156], Arafura Sea, Australia [157]

Hong Kong [133]

Hong Kong [158]

Cephalopholis argus (Blue-spotted grouper, Roi)

Cephalopholis miniata (Coral cod/Coral grouper)

Epinephelus coioides (Orange-spotted grouper)

Epinephelus lanceolatus (Giant grouper)

GrouperC

CTX – positive [155]

New Caledonia, French Polynesia [155]

Tridacna sp. (Giant Clam)

Giant ClamH

Conus spp. (Cone snails)

CTX – positive [154]

CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Parupeneus insularis (Twosaddle goatfish)

Hawaii [154]

CTX – positive [144]

St. Barthelemy, Caribbean Sea [144]

Mulloidichthys martinicus (Yellow goatfish)

GastropodC

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

MBA [158]

MBA [133]

HPLC/MS [157], MBA [157], N2A [60, 156]

Cat BA [149], MQBA [149], MBA [149], ELISA [129], N2A [129], RBA [54]

N2A [155], RBA [155]

MBA [155], N2A [155], RBA [155]

Ciguatect® [154]

RBA [54]

TLC [144], MBA [144]

Method of detection

292 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

MBA [110] Cat BA [149], MQBA [149], MBA [133, 149, 161], RBA [54]

CTX – positive [126] CTX – positive [144] CTX – positive [159] C-CTX-1 [60], C-CTX-2 [60] CTX-1 [160] CTX – positive [159] CTX – positive [148] CTX – positive [133] CTX – positive [110] CTX – positive [54, 133, 149, 161]

St. Thomas, Carribean Sea [126]

St. Barthelemy, Caribbean Sea [144]

Baja California, Mexico [159]

Key Largo, Florida, USA [60]

Baja California, Mexico [160]

Baja California, Mexico [159]

Guadeloupe and St. Barthelemy, Caribbean Sea [148]

Hong Kong [133]

Hong Kong [110]

French Polynesia, Tubuai (Australes) [54], Hong Kong [133], Tahiti [161], French Polynesia [149]

Epinephelus mystacinus (Misty grouper)

Epinephelus morio (Red grouper)

Epinephelus sp.

Mycteroperca bonaci (Black grouper)

Mycteroperca prionura (Sawtail grouper)

Mycteroperca sp.

Mycteroperca venenosa (Yellowfin grouper)

Plectropomus areolatus (Squaretail coral grouper)

Plectropomus laevis (Blacksaddled coral grouper)

Plectropomus leopardus (Coral trout/leopard coral grouper)

MBA [133]

Chick BA [148]

MBA [159]

HPLC/MS [160], MBA [160]

LCMS/MS [60], N2A [60]

MBA [159]

TLC [144], MBA [144]

BSBA [126], MGBA [126]

Cat BA [149], MQBA [149], MBA [149]

CTX – positive [149]

French Polynesia [149]

Epinephelus microdon (Marble grouper)

Method of detection

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

9 Gambierdiscus, the cause of ciguatera fish poisoning |

293

Caranx ignobilis (Giant trevally (ulua))

Jacks and

French Polynesia, Tubuai (Australes) [54], St. Barthelemy, Caribbean Sea [144]

Hawaii [132]

Bodianus sp.

CTX – positive [54, 131]

CTX – positive [132]

CTX – positive [144]

St. Barthelemy, Caribbean Sea [144]

Bodianus rufus (Spanish hogfish)

ScadsC

CTX – positive [131]

Hawaii [131]

Bodianus bilunulatus (Tarry hogfish (a’awa))

HogfishC

Pomadasys maculatus (Blotched javelin)

CTX-1 [162], CTX-2 [162], CTX-3 [162]

CTX – positive [110]

Hong Kong [110]

Variola albimarginata (Lyretail)

Platypus Bay, Queensland, Australia [162]

C-CTX-1 [70, 163], C-CTX-2 and isomers [70, 163], CTX congeners, other compounds [70, 163]

French West Indies [70, 163]

Serranidae

GruntC

CTX-1 [162], CTX-2 [162], CTX-3 [162]

Great Barrier Reef, Australia [162]

Plectropomus spp. (Coral trout)

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

MBA [131], S-EIA [131], RBA [54]

SPIA [132]

TLC [144], MBA [144]

MBA [131], S-EIA [131]

HPLC/MS [162], MBA [162]

MBA [110]

MBA [70, 163]

HPLC/MS [162], MBA [162]

Method of detection

294 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

CTX – positive [54]

French Polynesia, Tubuai (Australes) [54]

Hawaii [131, 132]

Caranx papuensis (Brassy trevally)

Caranx sp. (Trevally (ulua, papio))

LCMS/MS [60], TLC [144], Chick BA [148], MBA [144], N2A [60] HPLC/MS [162], TLC [135], MBA [135, 150, 162]

C-CTX-1 [60], C-CTX-2 [60]

CTX-1 [162], CTX-2 [162], CTX-3 [162]

Hervey Bay, Queensland, Australia [150], Hervey Bay, Queensland, Australia [135]

Scomberomorus commerson (Spanish mackerel)

SPIA [164]

Florida, USA [60], St. Barthelemy, Caribbean Sea [144, 148], Guadeloupe [148]

CTX – positive [164]

MBA [131], S-EIA [131, 132] SPIA [131, 132]

RBA [54]

Cat BA [149], MQBA [149], MBA [149], RBA [54]

Scomberomorus cavalla (King mackerel “Coronado” (Kingfish))

MackeralO

Cnidaria sp.

American Samoa [164]

CTX – positive [54, 149]

Nuku Hiva (Marquesas) [54], French Polynesia [149]

Caranx melampygus (Bluefin trevally)

JellyfishO

C-CTX-1 and isomers, CTX congeners [70, 163]

French West Indies [70, 163]

Caranx lugubris (Black jack)

CTX – positive [131, 132]

HPLC/MS [68–70, 163], BSBA [126], Cat BA [123], MGBA [126], MBA [68, 70, 163]

12 CTXs (inc C-CTX-1, C-CTX-1a, C-CTX-2) [70, 163], C-CTX-1 [68, 69] and C-CTX-2 [68, 69]

French West Indies [70, 163], St. Barthelemy, Caribbean Sea [68, 69], The Bahamas [123], St. Thomas, Carribean Sea [126]

Caranx latus (Horse-eye jack)

MBA [70, 163]

Method of detection

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

9 Gambierdiscus, the cause of ciguatera fish poisoning |

295

CTX – positive [54] CTX – positive [54] CTX – positive [54] CTX – positive [54] CTX-4A [101] CTX – positive [149]

French Polynesia, Tubuai (Australes) [54]

French Polynesia, Tubuai (Australes) [54]

French Polynesia, Tubuai (Australes) [54]

French Polynesia, Tubuai (Australes) [54]

French Polynesia [101], Tahiti [161], French Polynesia [149]

French Polynesia [149]

Chlorurus microrhinos (Steephead parrotfish)

Scarus altipinnis (Filament-finned parrotfish)

Scarus ghobban (Blue-barred parrotfish)

Scarus gibbus (Heavy beak parrotfish)

Scarus jonesi

CTX-3C [165]

Chlorurus frontalis (Pacific slopehead parrotfish)

ParrotfishH

Oplegnathus punctatus (Spotted knifejaw)

Miyazaki, Japan [165]

CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Liza vaigiensis (Thinlip grey mullet)

KnifejawO

CTX – positive [54]

CTX (if detected)

Nuku Hiva (Marquesas) [54], French Polynesia [149]

Source

Crenimugil crenilabis (Fringelip mullet)

MulletO

Latin name (Common name)

Tab. 9.4 (continued)

Cat BA [149], MQBA [149], MBA [149]

HPLC/HNMR [101], MQBA, MBA [101, 149, 161]

RBA [54]

RBA [54]

RBA [54]

RBA [54]

HPLC/MS [165]

RBA [54]

MQBA [149], MBA [149], RBA [54]

Method of detection

296 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

CTX – positive [132]

CTX – positive [149]

French Polynesia [149]

Aprion virescens (Bluge green snapper)

P-CTX-3C [171]

CTX – positive [169, 170]

CTX – positive [54]

CTX – positive [167, 168]

Hawaii [132]

Hawaii [171]

Hawaii [169, 170]

French Polynesia, Tubuai (Australes) [54], Nuku Hiva (Marquesas) [54]

Chile [167, 168]

CTX – positive [166]

Aphareus furca (Black forktail snapper (wahanui))

SnapperC

Monachus schauinslandi (Hawaiian monk seal)

SealC

Holothuria spp.

Sea cucumberH

Kyphosus cinerascens (Blue sea chub)

Sea

chubO

Farmed salmon

SalmonO

Siganus rivulatus (Marbled spinefoot)

Eastern Mediterranean [166]

CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Scarus rubroviolaceus (Ember parrotfish)

RabbitfishH

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

Cat BA [149], MQBA [149], MBA [149]

S-EIA [132], SPIA [132]

LCMS/MS [171], N2A [171]

Ciguatect® [169, 170]

RBA [54]

SPIA [167, 168]

Cigua-check® [166]

RBA [54]

Method of detection

9 Gambierdiscus, the cause of ciguatera fish poisoning |

297

HPLC/MS [72, 73], Cat BA [149], MGBA [72, 73], MQBA [149], MBA [72, 73, 149], RBA [54]

I-CTX-1 [72, 73], CTX-1B [165]

CTX – positive [127] CTX – positive [54, 149] C-CTX-1 and isomers [70, 163], CTX congeners [70, 163] CTX – positive [131] CTX – positive [54] I-CTX [72, 73], I-CTX-2 [72, 73], I-CTX-3 [72, 73], I-CTX-4 [72, 73]

Republic of Mauritius [72, 73], Minamitorishima (Marcus) Island, Japan [165], French Polynesia, Tubuai (Australes) [54] ,Nuku Hiva (Marquesas) [54], Hawaii [132], French Polynesia [149]

St. Croix, US Virgin Islands [127]

Nuku Hiva (Marquesas) [54], French Polynesia [149]

French West Indies [70, 163]

Hawaii [131]

Nuku Hiva (Marquesas) [54]

Republic of Mauritius (Nazareth, Saya de Malha, Soudan) [72, 73]

Lutjanus bohar (Two spot red snapper (red bass))

Lutjanus buccanella (Blackfin snapper)

Lutjanus gibbus (Humpback red snapper)

Lutjanus griseus (Grey snapper)

Lutjanus kasmira (Bluestripe snapper (taape))

Lutjanus monostigma (One-spot Snapper)

Lutjanus sebae (Red emperor)

HPLC/MS [72, 73], HPLC/MS/RLB [72, 73], MGBA [72, 73], MBA [72, 73]

RBA [54]

MBA [131], S-EIA [131], SPIA [131]

MBA [70, 163]

MQBA [149], MBA [149], RBA [54]

TLC [127], MBA [127]

MBA [110]

CTX – positive [110]

Hong Kong [110]

Lutjanus argentimaculatus (Mangrove red snapper)

Method of detection

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

298 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

CTX – positive [131]

CTX – positive [131]

Hawaii [131]

Hawaii [131]

Acanthurus nigroris (Bluelined surgeonfish (maiko))

CTX – positive [154]

Acanthurus dussumieri (Dussumier’s surgeonfish (palani))

SurgeonfishH

Ophiocoma spp. (Ophiuroids (brittle stars))

Hawaii [154]

CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Sargocentron spiniferum (Sabre squirrelfish)

StarfishO

CTX – positive [131]

Hawaii [131]

Myripristis kuntee (Epaulette Soldierfish (squirrelfish))

Squirrelfish and SoldeirfishC

MBA [131], S-EIA [131]

MBA [131], S-EIA [131]

Ciguatect® [154]

RBA [54]

MBA [131], S-EIA [131], SPIA [131]

MBA [110]

CTX – positive [110]

Hong Kong [110]

Lutjanus stellatus (Star snapper)

HPLC/MS [165], BSBA [126], MGBA [126], MBA [172], S-EIA [132], SPIA [132], N2A [49]

CTX-1B [165]

Antigua [132], Okinawa, Japan [165], West Africa [49], Baja California, Mexico [172], St. Thomas, Carribean Sea [126]

Lutjanus spp. (Snapper)

Method of detection

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

9 Gambierdiscus, the cause of ciguatera fish poisoning |

299

CTX – positive [54] CTX – positive [54, 173]

Nuku Hiva (Marquesas) [54]

Nuku Hiva (Marquesas) [54], Tahiti [173]

Acanthurus xanthopterus (Yellowfin surgeonfish)

Ctenochaetus striatus (Striped Bristletooth)

CTX – positive [54] CTX – positive [54] CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Nuku Hiva (Marquesas) [54]

Naso brevirostris (Spotted unicornfish)

Naso hexacanthus (Sleek unicornfish)

CTX – positive [54]

Nuku Hiva (Marquesas) [54]

Nuku Hiva (Marquesas) [54], French Polynesia [149]

Naso brachycentron (Humpback unicornfish)

UnicornfishC

Gymnosarda unicolor (Dogtooth tuna)

TunaC

Malacanthus plumieri (Sand tilefish)

CTX – positive [144]

CTX – positive [132]

Hawaii [132]

Acanthurus sp.

St. Barthelemy, Caribbean Sea [144]

CTX – positive [131]

Hawaii [131]

Acanthurus olivaceus (Orangeband surgeonfish (naenae))

TilefishC

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

RBA [54]

RBA [54]

RBA [54]

Cat BA [149], MQBA [149], MBA [149], RBA [54]

TLC [144], MBA [144]

RBA [54]

RBA [54]

S-EIA [132], SPIA [132]

MBA [131], S-EIA [131]

Method of detection

300 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

CTX – positive [54, 149] CTX – positive [54]

Nuku Hiva [54, 149] (Marquesas)

Nuku Hiva (Marquesas) [54]

Naso lituratus (Orangespine unicornfish)

Naso unicornis (Bluespine unicornfish)

Baja California, Mexico [159]

Semicossyphus sp.

MBA [159]

RBA [54]

Cat BA [149], MQBA [149], MBA [133, 149]

RBA [54]

Cat BA [149], MQBA [149], MBA [149], RBA [54]

Method of detection

The abbreviations are: LC-MS/MS: liquid chromatography tandem mass spectrometry, UPLC/MS: ultra performance liquid chromatography/mass spectrometry, HPLC/MS: high performance liquid chromatography/mass spectrometry, HPLC/HNMR: high performance liquid chromatography/H nuclear magnetic resonance, HPLC/MS/RLB: high performance liquid chromatography/mass spectrometry/radio ligand binding, TLC: thin layer chromatography, BSBA: brine shrimp bioassay, DLBA: diptera larvae bioassay, MGBA: mongoose bioassay, MQBA: mosquito bioassay, MBA: mouse bioassay, SEIA: stick enzyme immunoassay, SPIA: solid phase immunoassay, RIA: radioimmunoassay, ELISA: enzyme-linked immunosorbent assay, N2A: neuroblastoma cytotoxicity assays, RBA: receptor-binding assay, MA: membrane assay, BA: bioassay.

Typical feeding behavior: C: carnivore, H: herbivore, O: omnivore.

CTX – positive [54]

French Polynesia, Tubuai (Australes) [54]

Coris aygula (Clown coris) CTX – positive [159]

CTX – positive [133, 149]

French Polynesia [149], Hong Kong [133]

Cheilinus undulatus (Humphead Wrasse)

WrasseC

CTX (if detected)

Source

Latin name (Common name)

Tab. 9.4 (continued)

9 Gambierdiscus, the cause of ciguatera fish poisoning |

301

302 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

The CTX-positive cases in Tab. 9.4 are predominantly concerned with the mid-latitude tropical and sub-tropical zones. This is fitting with the distribution of Gambierdiscus as described in Tab. 9.2. However, CFP has also been reported in non-endemic areas because of an increase in seafood imports [114, 117]. While the majority of studies have focused on reef fish, toxin accumulation has been observed in eels, sea cucumbers, starfish, seals and jellyfish (see Tab. 9.4 and references therein). Sharks have also been suspected of causing CFP following outbreaks of human illness, remnant samples for testing were unavailable [118, 119]. Further studies are required to address the deficit in information for species other than fish and to identify potential toxin vectors in coastal systems. For the most part, CFP studies have focused on CTX rather than MTX. The MBA has been used previously to test for MTX, with positive results in Ctenochaetus striatus (striped bristletooth) (Bagnis et al., 1986). A gap in our existing knowledge is whether the presence of MTX in small (herbivorous) fish species is transferred up the food chain to larger carnivorous species. Often, in small island nations, native fishermen are aware of ciguatera prone zones and avoid certain fish species. Such knowledge certainly has its merits; however, a study by [54] in French Polynesia demonstrated the presence of CTXs in fish species that were considered safe to eat by locals. Experimentally, CTX toxin profiles and structures have been determined by chromatographic techniques (HPLC, UPLC and LC-MS), accompanied by nuclear magnetic resonance (NMR) [62, 63, 101, 120] and radio ligand binding (RLB) [72, 73]. However, these methods are not commonplace or practical for routine testing, as they are costly and require special expertise. Confirmation of toxin by UPLC/HPLC followed by LC-MS involves the isolation and fractionation of the various CTX compounds and their known molecular weights (see Tab. 9.3). Although a rapid method for sample analysis has been proposed [121], acquiring purified CTX standards is problematic due to the limited supply of natural CTX compounds [48]; though artificial synthesis of CTX is possible [122], it is highly complex. Without a consistent source of reference material, absolute quantification of CTXs and their congeners is hard to achieve. In addition, technical issues such as co-eluting peaks of similar compounds and inhibiting/promoting matrix effects remain unresolved. Several biological assays have been developed for the detection of ciguateric fish. These have included the use of chickens (Pottier et al., 2000), cats [123], mongooses [124], diptera larva [125], brine shrimp [126] and mosquitos (Bagnis et al., 1987). However, each assay has its own constraints and limitations, largely relating to toxin specificity and quantification, but also due to inefficiencies and ethical considerations (summarized in de Fouw, 2001, [109]). While the MBA by intraperitoneal injection does not provide a linear dose-response relationship with CTX toxicity [127], it remains the most widely used biological assay (see Tab. 9.4). Numerous biochemical assays have been proposed as alternatives to biological assays for testing seafood. The development of a radioimmunoassay [124] progressed to a cheaper alternative enzyme-linked immunosorbent assay (ELISA) with higher throughput [128]. The ELISA test has recently shown promising correlations with

9 Gambierdiscus, the cause of ciguatera fish poisoning

|

303

biological assays [129, 130]. Stick enzyme immunoassay (SEIA) [131] and solid phase immunoassay (SPIA) [132] tests have led to the development of commercial kits (i.e. Cigua-check® and Ciguatect® ). However, these products have yielded a large number of false positive and false negative results [133] and the Cigua-check® test is no longer being manufactured. Other assays utilized for screening CTXs in fish are the sodium channel binding assay (N2A) [60] and RBA [54, 134]. Both of these assays have shown promising results and have been recommended by the European Food Standard Association (EFSA, 2010). These assays cannot quantify specific congeners of CTXs and MTXs. This can only be achieved via further development and validation via LC-MS analysis, and there is an urgent need to do so. The progress has been disadvantaged by the lack of available purified standards (Guzman-Perez and Park, 2000). Other challenges are the presence of more than one type of CTXs, (see Tab. 9.3) being present in fish specimens [68, 135].

9.7 Conclusion Since its recognition as the source of CFP, major advances have been made in the study of Gambierdiscus species.Concurrently, new questions and challenges have also been raised. Here, we outline the major areas needing research efforts to significantly advance our understanding of the causes of the production of toxins leading to CFP: 1. It is highly likely that species of Gambierdiscus vary in their toxicity, whereas intraspecific toxin production appears to be more consistent. Exact Gambierdiscus species identifications in CFP affected areas around the world are therefore required. New molecular and taxonomic tools to identify Gambierdiscus species accurately and simply will therefore be required. 2. Toxin profiles of Gambierdiscus strains and species are needed to identify the exact CTXs and MTXs produced. 3. The development and standardization of chromatographic techniques to accurately quantify different CTXs and MTXs are required. This involves a further characterization of already known and new congeners of CTXs and MTXs. 4. The development of commercially available CTX and MTX standards is very important, as it is one of the major hurdles that prevent further advancement of areas mentioned in the above two goals. 5. The elucidation of genes involved in biosynthesis of CTX and MTX in Gambierdiscus species will allow for an increased understanding of the causes and triggers of toxin production and the potential for the development of novel CFP monitoring tools.

304 | Gurjeet S. Kohli, Hazel Farrell, and Shauna A. Murray

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[140] Mohammad-Noor N, Daugbjerg N, Moestrup Ø, Anton A. Marine epibenthic dinoflagellates from Malaysia – a study of live cultures and preserved samples based on light and scanning electron microscopy. Nord J Bot 2007;24. [141] Jeong HJ, Lim AS, Jang SH, et al. First report of the epiphytic dinoflagellate Gambierdiscus caribaeus in the temperate waters off Jeju Island, Korea: morphology and molecular characterization. J Eukaryot Microbiol 2012;59:637–50. [142] Caillaud A, Eixarch H, de la Iglesia P, et al. Towards the standardisation of the neuroblastoma (neuro-2a) cell-based assay for ciguatoxin-like toxicity detection in fish: application to fish caught in the Canary Islands. Food Addit Contam 2012;29:1000–10. [143] Otero P, Perez S, Alfonso A, et al. First toxin profile of ciguateric fish in Madeira Arquipelago (Europe). Ann Chem 2010;82:6032–9. [144] Vernoux JP, Elandaloussi SA. Heterogeneity of ciguatoxins extracted from fish caught at the coast of the french-antilles. Biochimie 1986;68:287–91. [145] Boada LD, Zumbado M, Luzardo OP, et al. Ciguatera fish poisoning on the West Africa Coast: an emerging risk in the Canary Islands (Spain). Toxicon 2010;56:1516–9. [146] Pérez-Arellano J-L, Luzardo OP, Brito AP, et al. Ciguatera fish poisoning, Canary Islands. Emerging Infect Dis 2005;11:1981. [147] O’Toole AC, Bottein MYD, Danylchuk AJ, Ramsdell JS, Cooke SJ. Linking ciguatera poisoning to spatial ecology of fish: A novel approach to examining the distribution of biotoxin levels in the great barracuda by combining non-lethal blood sampling and biotelemetry. Sci Total Environ 2012;427:98–105. [148] Pottier I, Vernoux J-P, Lewis RJ. Ciguatera fish poisoning in the Caribbean islands and Western Atlantic. In: Ware GW, Nigg HN. [eds.] Reviews of environmental contamination and toxicology. Tuscon, Arizona; Springer: 2001. p. 99–141. [149] Bagnis R, Barsinas M, Prieur C, Pompon A, Chungue E, Legrand A. The use of the mosquito bioassay for determining the toxicity to man of ciguateric fish. Biol Bull 1987;172:137–43. [150] Lewis RJ, Endean R. Ciguatoxin from the flesh and viscera of the barracuda, Sphyraena jello. Toxicon 1984;22:805–10. [151] Hung YM, Hung SY, Chou KJ, et al. Short report: Persistent bradycardia caused by ciguatoxin poisoning after barracuda fish eggs ingestion in southern Taiwan. Am J Trop Med Hyg 2005; 73:1026–7. [152] Lewis RJ, Jones A. Characterization of ciguatoxins and ciguatoxin congeners present in ciguateric fish by gradient reversed-phase high-performance liquid chromatography/mass spectrometry. Toxicon 1997;35:159–68. [153] Scheuer PJ, Takahashi W, Tsutsumi J, Yoshida T. Ciguatoxin: isolation and chemical nature. Science 1967;155:1267–8. [154] Park DL, Ayala CE, Guzman-Perez SE, Lopez-Garcia R, Trujillo S. Microbial toxins in foods: algal, fungal, and bacterial. Food Toxicology. New York; CRC Press: 2001. p. 93–135. [155] Laurent D, Kerbrat A, Darius H, et al. Ciguatera Shellfish Poisoning, a new ecotoxicological phenomenon from cyanobacteria to human via giant clams. Food chain: new research 2012: 1–44. [156] Arnett MV, Lim JT. Ciguatera fish poisoning – impact for the military health care provider. Mil Med 2007;172:1012–5. [157] Lucas RE, Lewis RJ, Taylor JM. Pacific ciguatoxin-1 associated with a large common-source outbreak of ciguatera in East Arnhem Land, Australia. Nat Toxins 1997;5:136–40. [158] Wong CK, Hung P, Lee KLH, Kam KM. Solid-phase extraction clean-up of ciguatoxincontaminated coral fish extracts for use in the mouse bioassay. Food Addit Contam 2009; 26:236–47.

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[159] Lechuga Solid-phas Sierraa Solián AP. Documented case of ciguatera on the Mexican Pacific coast. Nat Toxins 1995;3:415–8. [160] Sierra-Beltran AP, Cruz A, Nunez E, Del Villar LM, Cerecero J, Ochoa JL In: An overview of the marine food poisoning in Mexico, 12th World Congress on Animal, Plant and Microbial Toxins, Cuernavaca, Mexico, Sep 21-26, 1997. Cuernavaca, Mexico; 1997. p. 1493–1502. [161] Pompon A, Bagnis R. Ciguatera – a rapid procedure for extraction of ciguatoxin. Toxicon 1984;22:479–82. [162] Lewis RJ, Sellin M. Multiple ciguatoxins in the flesh of fish. Toxicon 1992;30:915–9. [163] Pottier I, Vernoux J, Jones A, Lewis R. Analysis of toxin profiles in three different fish species causing ciguatera fish poisoning in Guadeloupe, French West Indies. Food Addit Contam 2002;19:1034–42. [164] Zlotnick BA, Hintz S, Park DL, Auerbach PS. Ciguatera poisoning after ingestion of imported jellyfish – diagnostic application of serum immunoassay. Wilderness Environ Med 1995; 6:288–94. [165] Yogi K, Oshiro N, Inafuku Y, Hirama M, Yasumoto T. Detailed LC-MS/MS analysis of Ciguatoxins revealing distinct regional and species characteristics in fish and causative alga from the Pacific. Ann Chem 2011;83:8886–91. [166] Bentur Y, Spanier E. Ciguatoxin-like substances in edible fish on the eastern Mediterranean Clin Toxicol 2007;45:695–700. [167] Ebesu JSM, Nagai H, Hokama Y. The first reported case of human ciguatera possibly due to a farm-cultured salmon. Toxicon 1994;32:1282–6. [168] Dinubile MJ, Hokama Y. The ciguatera poisoning syndrome from farm-raised salmon. Ann Intern Med 1995;122:113–4. [169] Park DL. Seafood safety monitoring programme for ciguatera: assessing aquatic product safety. Proc Gulf Caribb Fish Inst 1999;45:270–89. [170] Dalzell P. Management of ciguatera fish poisoning in the South Pacific. Memoirs of the Queensland Museum. Brisbane 1994;34:471–9. [171] Bottein M-YD, Kashinsky L, Wang Z, Littnan C, Ramsdell JS. Identification of ciguatoxins in Hawaiian monk seals Monachus schauinslandi from the Northwestern and Main Hawaiian Islands. Environ Sci Technol 2011;45:5403–9. [172] Parrilla-Cerrillo MC, Vázquez-Castellanos JL, Saldate-Castañeda EO, Nava-Fernández LM. Outbreaks of food poisonings of microbial and parasitic origins. Brotes de toxiinfecciones alimentarias de origen microbiano y parasitario 1993;35:456–63. [173] Bagnis R, Bennett J, Prieur C, Legrand AM. In: The dynamics of three toxic benthic dinoflagellates and the toxicity of ciguateric surgeonfish in French Polynesia, Third International Conference on Toxinc Dinoflagellates, St. Andrews, New Brunswick, Canada, June 8–12, 1985. Andersen D, White A, Baden DG. [eds.] St. Andrews, New Brunswick, Canada; Elsevier: 1985. p. 177–82.

Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

10 Control and management of Harmful Algal Blooms 10.1 Introduction Everyone would agree that the best cure is prevention. This is certainly applicable in the case of cyanobacterial blooms and their tremendous and widespread impact on human and environmental health, natural and man-made assets, as well as overall ecosystem services. There have even been suggestions and also some evidence that cyanobacteria blooms may have a negative impact on the housing markets in some parts of the world. Certainly, common sense would dictate that the level of appreciation and the livability of a neighborhood may be positively influenced by the presence of a nice lake (natural or artificial); however, the opposite is true when that lake is affected by water quality issues including odor and toxins from the development of algal blooms. Prevention of cyanobacterial blooms has been at the heart of the nutrient management strategies around the world and has resulted in many success stories. However, blooms still occur and they also occur in assets that the public are not necessarily exposed to, such as drinking water reservoirs, wastewater stabilization ponds and retention dams, in addition to natural systems such as lakes and rivers. When prevention fails or less than satisfactory results are achieved, we need to tackle the problem through direct mitigation technologies that aim at reducing or eliminating the risk generated by the presence of cyanobacteria and their toxins. In this chapter, we will discuss a broad range of mitigation approaches that have been applied with various degrees of success, and we also explore the opportunities for future development of innovative solutions for this important problem.

10.2 Global water crisis Humans utilize water in agriculture, industry, the household and for recreation [1]. Water also provides incidental ecosystem services to humans and other organisms, such as habitats, climate control and media for nutrient cycling [2]. It is a natural resource essential to life on earth, and adequate management of water is required to maintain these vital services to the global population and the environment. Water resources are generally not well managed, and as a result, water quality remains of significant concern. Waterways are continually polluted through the addition of nutrients and heavy metals, high-quality water is wasted on activities where it is not suited to the purpose, and water is not equally distributed between people of all nations and social status. These anthropogenic impacts on water quality and quantity have led to a crisis that is affecting the entire biosphere of the earth [1, 3, 4]. This crisis

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is increasing with the concurrent rise in global population and climate change [4, 5]. In order to combat this, water resource management must be adaptive, flexible and engage with stakeholders at multiple levels, from the local to the global. One of the most significant anthropogenic impacts on waterways is eutrophication. Eutrophication occurs when high nutrient loads enter a water body, often as a result of agricultural and industrial processes. This increase in nutrients, particularly nitrates and phosphates, can detrimentally affect ecosystems and reduce the quality of water for reuse purposes [6, 7]. Eutrophication is often a precursor to the occurrence of harmful algal blooms, which commonly contain toxin-producing cyanobacterial species and are a threat to human and environmental health [8].

10.3 Cyanobacteria and cyanotoxins Cyanobacteria are prokaryotic phytoplanktons that occur in fresh, brackish and salt water systems throughout the world [9, 10]. Species of cyanobacteria differ in their morphology and may exist as single cells, colonies and filaments [11]. When cells aggregate, they form dense cyanobacterial blooms, a potential threat to human and environmental health. Cyanobacteria generally dominate in reservoirs containing high nutrient loads and stagnant water, although cyanobacterial blooms do occasionally occur in oligotrophic systems and favor water temperatures between 15 and 30 °C [8]. Studies into the dynamics of cyanobacterial blooms predict that the expected increase in global temperature will result in increased surface water temperatures and thermal stratification, as well as changing meteorological patterns, possibly stimulating increased cyanobacterial growth rates [12–19]. It is likely that this will result in an increased frequency of cyanobacterial bloom events. Of particular concern to water utility managers are those cyanobacterial species that form blooms in freshwater reservoirs that are used for drinking, recreation and irrigation. Cyanobacterial blooms have several detrimental environmental effects. Blooms often proliferate in the surface layer of stratified reservoirs, shading organisms below, which can result in the death of pelagic and benthic organisms [20–23]. When blooms collapse, the release of organic cell matter to the water column increases the system’s oxygen demand. The concentration of dissolved oxygen is lowered due to its consumption in reactions to degrade organic and inorganic compounds; this results in mass deaths of fish and other aquatic organisms [10, 24]. Such deaths are often observed by the general public and receive considerable media attention. Many species of cyanobacteria also produce toxins. Cyanobacterial toxins (cyanotoxins) vary in their toxicity to humans and animals, and include hepatotoxins, dermatoxins, cytotoxins, neurotoxins and lipopolysaccharides. Cyanotoxins can induce both acute and chronic effects, and can pose a risk to both humans and ecological systems [25–32].

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The most common routes of human contact with cyanotoxins are through the contamination of drinking water, the recreational use of lakes and rivers containing cyanobacteria and via the ingestion of blue-green algal supplements [33–37]. Organisms within the environment are often harmed by direct exposure to cyanotoxins or through bioaccumulation [38–48]. Bioaccumulation can lead to the magnification of cyanotoxins throughout food webs, potentially altering growth patterns, grazing behavior and development, and leading to significant health risks for organisms, including humans, that predate species which have bio-accumulated cyanotoxins [29, 41, 49, 50]. The shading of underlying organisms, reduction of dissolved oxygen and bioaccumulation of cyanotoxins can lead to shifts in ecological assemblages and potentially ecosystem collapse, as well as significant threats to human health. As such, it is imperative that the risks of cyanobacterial blooms in various freshwater bodies are assessed and mitigated so that they can be appropriately managed to avoid detrimental effects.

10.4 Cyanobacterial prevention and mitigation Many techniques for cyanobacterial bloom prevention and mitigation have been investigated (Tab. 10.1 and Tab. 10.2). Some have been applied directly in reservoir management, while others have been trialed only under laboratory conditions. The success of preventative and mitigation techniques depends upon the underlying conditions present, and the characteristics of individual water bodies must be considered when determining the most appropriate management strategies to apply. Prevention of cyanobacterial blooms has been achieved with varying success through techniques including nutrient reduction, artificial destratification, macrophyte establishment, predation, the addition of allelopathic chemicals, ultraviolet radiation (UVR) and ultrasonication (Tab. 10.1). Nutrient reduction and destratification have shown reasonable success in large reservoirs, though most success has been where nutrient inputs can be significantly reduced and reservoirs are relatively deep. Despite preventative attempts, often cyanobacterial blooms still occur. It is therefore imperative that mitigation measures for controlling blooms are investigated. Many such methods have been trialed in both the laboratory and field, with varying success (Tab. 10.2). It is common practice to remove cyanobacteria using copper sulfate, chlorine or coagulants and flocculants [109], although the dynamics of the removal of cyanobacteria from wastewater by such methods has not been thoroughly investigated. These cyanobacterial removal techniques currently practiced on a large scale may be environmentally damaging and ineffective for the removal of cyanotoxins [93, 109–112]. Several of the removal methods used in drinking water treatment are highly successful where cyanobacterial and cyanotoxin concentrations are low and the water will not

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be released to the environment, but are often prohibitively expensive for use in highly eutrophic systems and generally less effective in reservoirs containing high concentrations of organic matter [112, 113] (Tab. 10.2). Tab. 10.1: Commonly used prevention strategies for cyanobacterial blooms. Specific comments for use

References

Allelopathic chemicals

Chemicals generally secreted by decaying organic matter. Addition of organic matter increases WSP sludge production. May be unanticipated effects on non-target organisms.

[51–64]

Destratification

Can promote growth of non-buoyant phytoplankton over cyanobacteria. Often ineffective in shallow and highly eutrophic water bodies. Generally requires electrical connection on-site.

[65–71]

Macrophytes

Interfere with WSP processes. Provide breeding grounds for mosquitos and other disease vectors.

[72–81]

Nutrient reduction

50 % of phosphorus in wastewater is from human waste and cannot be reduced. Likely that WSPs will be high in nutrients regardless of reduction measures. N:P ratio may be more important than actual phosphorus and nitrogen concentrations.

[8, 71, 82–85]

Predation

May alter WSP ecology, particularly if zooplankton are added and preferentially consume non-target phytoplankton. Consumption of cyanotoxins may result in the death of predators.

[86–91]

Ultrasonication

Only tested at reduced scales. May not be appropriate for full-scale WSPs.

[92–103]

Ultraviolet radiation

Cells in WSP are likely adapted to high UVR doses. May be practical in association with other treatment methods. Can only be used at pond inlets and outlets.

[104–108]

✓

Coagulation and flocculation

Increases sludge loading. Generally does not affect membrane integrity, so cyanotoxins are not released. If flocs are not removed, cyanotoxins accumulate in sludge. Must consider flow environment of WSP. Increases concentration of aluminum in the environment.

Ineffective at removing microcystins at pH > 8. Phytoplankton cells, rather than cyanotoxins, may preferentially react with chlorinated compounds. Cyanotoxins may be released from cells more quickly than they can be degraded by chlorinated compounds in solution. Produces by-products dangerous to humans and the environment (e.g. trihalomethanes).

Detailed in Tab. 10.3.

✓

Biodegradation ✓

Other organic compounds compete for adsorption-sites. May increase sludge loading.

✓

Adsorption (activated carbon)

✓

Occurs naturally, but may be insufficient for complete cyanotoxin removal. Cyanotoxin variants adsorb differently. Adsorption decreases as pH increases. Often biodegradation is greater than adsorption when in contact with particles. If cyanotoxins are filtered through natural soil and adsorption is insufficient, this can endanger aquifers.

✓

Adsorption (naturally occurring particles)

Chlorine and chlorinated compounds

Specific comments for use

Toxin removal

Cyanobacterial removal

Tab. 10.2: Commonly used mitigation strategies for cyanobacterial blooms.

[111, 143–150]

[93, 106, 131–142]

Tab. 10.3

[124–130]

[114–123]

References

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✓

✓

✓

✓

Hydrogen peroxide

Ozone

Cost prohibitive – may preferentially react with other organic compounds. Cyanotoxin degradation decreases with increasing pH. Release of cellular organic compounds increases the ozone dose required. Requires electricity on-site for the production of ozone.

Effectiveness may be increased by presence of UVR and/or iron. Cyanotoxin degradation decreases with increasing pH. May release cyanotoxins to the dissolved state over hours/days. Often found ineffective on scales of minutes when not coupled with other methods. May affect microcystin synthesis within cells.

Impractical for water containing high suspended sediment loads. Most degradation in successful studies appears to be biological, except where nanofiltration is used.

✓

✓

Filtration and reverse osmosis

Specific comments for use Traditional method of cyanobacterial removal in WSPs. Releases cyanotoxins to the dissolved state, but does not subsequently degrade them. Increases copper concentration in the environment.

Toxin removal

✓

Cyanobacterial removal

Copper sulfate

Tab. 10.2 (continued)

[93, 106, 111– 113, 129, 132, 140, 180–182]

[129, 140, 141, 152, 164–179]

[157–163]

[110, 151–156]

References

318 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

✓ ✓

✓

Ultraviolet radiation

Occurs naturally, but can be enhanced by coupling with other removal methods. Can reduce cyanotoxins through photosensitized reaction with compounds including phycocyanin. Currently used to ensure the removal of coliforms at some WWTP outlets. Impractical alone where cell concentrations are high. Requires electricity on-site.

Only tested at reduced scales. May not be appropriate for full-scale WSPs. Requires electricity on-site.

Large amounts of catalyst are required – up to mg per l – so impractical in wastewater.

May alter WSP ecology, particularly if zooplankton are added and preferentially consume non-target phytoplankton. Addition of organic matter increases sludge loading.

Does not seem to be significantly affected by pH. Likely produces harmful by-products.

✓ ✓

Specific comments for use

Toxin removal

✓

✓

Cyanobacterial removal

Ultrasonication

Titanium dioxide

Predation and biomanipulation

Permanganate

Tab. 10.2 (continued)

[93, 105, 106, 165, 170, 174, 181, 189– 194]

[92–102]

[167, 186–188]

[51–54, 56, 58– 60, 86–90, 185]

[93, 106, 132, 136, 140, 141, 183, 184]

References

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319

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Effective techniques for the mitigation of both cyanobacteria and cyanotoxins must be determined by considering the underlying properties of the water system in question, including depth, pH values, concentrations of suspended solids, and dissolved organic and inorganic compounds [195, 196]. To be successful, any mitigation approach must reduce both cyanobacteria and cyanotoxins and pose no or negligible threat to ecosystems.

10.5 Cyanobacterial management The management of cyanobacteria in any freshwater resource must consider the entire cycle of water through catchments, water reservoirs, treatment plants and distribution systems [71]. Appropriate management of cyanobacterial blooms is imperative to reducing their negative impacts on human and ecological health, water treatment processes and income-generating activities, including tourism and property development. This is not simply about implementing prevention and mitigation strategies. It is also important to assess the potential risks associated with such blooms, so that they can be treated effectively and efficiently at the site of interest. Incorporating risk assessment into management will allow plans to be developed which minimize the costs of bloom mitigation and the potentially undesirable environmental effects of many cyanobacterial removal methods. Such plans should be developed for all freshwater resources that suffer from potentially toxic blooms. This will reduce or eliminate the undesirable consequences of cyanobacteria and cyanotoxins on both humans and ecosystems. A management plan should consist of the following actions required during four distinct time periods (Fig. 10.1): 1. Prior to bloom; 2. Hazardous bloom suspected; 3. Hazardous bloom identified; 4. Mitigation ineffective. Barrington et al. [197] offered a detailed approach for the appropriate monitoring regime to minimize the risk of undetected or undetectable incidents. Currently the authorities rely on some form of either public reporting of an incident or some form of monitoring which is usually a response to a visual inspection of the water system (Fig. 10.2). This form of risk assessment assumes that hazardous events will always be correctly identified. Although useful, such an approach may lead to overly cautious and cost-intensive behavior by following the precautionary principle, whereby potentially hazardous events are assumed to be dangerous regardless of their actual characteristics [198, 199]. If a monitoring methodology is overly precautionary, costs may be incurred by implementing unnecessary control measures to treat “false positive” results. This traditional method of assessment also fails to consider the associated risk should

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Risk of cyanobacterial bloom mitigated

Prevention

Routine monitoring and bloom identification

Contingency planning

Prior to bloom

Treatment of cyanobacterial bloom required

Cyanobacterial bloom suspected

Situation assessment

Bloom mitigation Assess the risk of remaining cyanobacterial bloom

Cyanobacterial bloom does not require treatment

Bloom isolation Treatment of cyanobacterial bloom insufficient

Contingency Invocation

Natural degradation

Hazardous bloom identified

Hazardous bloom suspected

Mitigation ineffective

Fig. 10.1: Management framework for the removal of toxic cyanobacteria from water bodies.

Tolerable risk for each Impact Potential Level 1.0 Low impact potential (HF = 1.00) Moderate impact potential (HF = 0.88) High impact potential (HF = 0.37)

0.8

Tolerable risk

0.6

0.4

0.2

0.0 0.0

0.2 0.4 0.6 Monitoring frequency (week–1)

0.8

1.0

Fig. 10.2: Tolerated/tolerable risk according to monitoring frequency for each impact potential level, for Swan Coastal Plain Lakes. The tolerated risk indicates the probability that a cyanobacterial bloom at or above each impact potential level will not be detected given it is occurring. HF = Hazard Frequency. For further details on the methodology and definitions, please refer to [203].

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a hazardous event not be identified by the monitoring regime (a “false negative” result). Since most monitoring methods are not of significant accuracy to identify every occurrence of a hazardous event, it is imperative that water managers are aware of the possibility that a dangerous event may occur which is not identified by the current monitoring methodology (Fig. 10.3). Water utilities and managers must optimize their monitoring regimes to reduce the occurrence of both “false positive” and “false negative” results, which will in turn reduce the risks and costs associated with hazardous events such as toxic cyanobacterial blooms. Tolerable risk for each impact potential level 1.0 Emu Jackadder Monger Bibra Blue gum Yangebup North Little rush

0.8

Tolerable risk

0.6

0.4

0.2

0.0 0.0

0.2 0.4 0.6 Monitoring frequency (week–1)

0.8

1.0

Fig. 10.3: Tolerated risk according to monitoring frequency for each impact potential level, at each of the investigated lakes of the Swan Coastal Plain. The tolerated risk indicated the probability that a cyanobacterial bloom at or above each impact potential level will not be detected given it is occurring. For further details on the methodology and definitions, please refer to [203].

A form of risk assessment that is used in medical diagnosis can be applied to environmental conditions and determines the probability that monitoring results will correctly identify hazardous situations (so long as the approximate frequency of the hazardous event is known). The development and application of this risk assessment methodology is outlined in Barrington et al. [197]. This assessment considers the relative probability of diagnostic tools returning “true positive”, “true negative”, “false positive” and “false negative” results [200]. This can assist in the development of monitoring programs and environmental decision-making [201, 202].

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10.6 Case study: The management of cyanobacteria in waste stabilization ponds Waste stabilization ponds (WSPs) are one form of freshwater body that have received minimal research with regards to the management of toxic cyanobacteria. Within such systems, the health and environmental risks associated with cyanobacterial blooms are coupled with the negative effects of cyanobacteria upon wastewater treatment processes, which may result in further indirect health, ecological and economic concerns. Cyanobacterial blooms are a serious problem in these systems, where they result in substantial increases to operational and maintenance costs of these assets. Hence the timely management of blooms in such reservoirs is essential. Water containing human excreta has been treated since the link between sewage and human health was first recognized. There are many methods of wastewater treatment currently utilized, but the most commonly used process throughout both the developed and developing worlds consists of systems of WSPs [196, 204]. Such wastewater treatment plants (WWTPs) are generally utilized in rural and remote areas (Fig. 10.4), but plants servicing upwards of one million people have shown success where the land is available and reasonably priced [204].

Fig. 10.4: Examples of waste stabilization ponds (WSPs) in Australia.

WSPs are a simple, highly efficient, low-cost, low-maintenance and robust process for treating wastewater [204–206]. In WSPs, wastewater constituents are removed by sedimentation or transformed by biological and chemical processes, and a sludge layer forms due to the sedimentation of influent suspended solids, algae, and bacteria [205, 207]. In addition, WSPs are more efficient at removing pathogens than the electrochemical methods utilized in most urban WWTPs [195, 208]. After coarse screening to remove large objects, wastewater enters an initial deep WSP where sedimentation removes settlable particles including helminth eggs and protozoan cysts. These ponds are likely anaerobic due to the high biological oxygen demand (BOD) loading, and such oxygen-depleted conditions result in the significant reduction of BOD. Following sedimentation and anaerobic processing, wastewater enters facultative WSPs, which primarily remove pollutants through algal-bacterial mutualism. The photosynthetic algae present in these WSPs produce oxygen, which is then consumed by the heterotrophic bacteria that degrade any remaining organic and

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inorganic compounds (Fig. 10.5) [195]. This further reduces the BOD of the water, such that it will not consume large amounts of dissolved oxygen when discharged. In the final WSP of the WWTP, referred to as the maturation or polishing pond, the majority of suspended sediments and pollutants have been reduced to acceptable levels for release. The primary function of the maturation pond is to kill dangerous wastewater organisms, including coliform bacteria and viruses, by the presence of natural radiation, high pH values and adsorption to settlable solids (Fig. 10.6) [195, 209]. In an ideal WWTP system, passage through these multiple WSPs will have decreased suspended sediment, organic and inorganic compounds, and dangerous wastewater organisms Light New cells Algae

CO2 NH4+ PO42–

O2

Bacteria New cells

Organic matter

Suspended solids ↓ BOD ↓ Pathogens ↓

Fig. 10.5: Algal-bacterial mutualism in waste stabilization ponds (adapted from [195]).

SUNLIGHT Rapid photosynthesis High DO pH > 9

Photo-oxidation

FAECAL BACTERIAL DIE-OFF

Increased pond temperature

Fig. 10.6: Conceptual mechanisms for faecal-bacterial die-off in facultative and maturation waste stabilization ponds (reproduced from [195]).

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to levels which will not harm humans or animals upon their release [195, 196]. Treated water is generally discharged to on-site evaporation, the environmental flow or human reuse [210]. Cyanobacteria have been recorded in WSPs throughout the world (e.g. [211–216], Fig. 10.7). Cyanobacterial blooms increase the sludge and suspended solids loadings of WSPs and change their ecology. By altering WSP ecology, cyanobacteria inhibit the natural processes of water purification anticipated by design engineers, particularly algal-bacterial mutualism, and this can result in the discharge of inadequately treated wastewater effluent.

Fig. 10.7: Cyanobacterial bloom at in WSP in central Western Australia (Photos: [217, 218]).

Cyanobacteria and cyanotoxins impact upon WSP ecology through physical, chemical and biological mechanisms [219]. The ability of cyanobacteria to regulate their buoyancy gives them a competitive advantage over other phytoplankton, forming dense blooms and surface scums that shade the organisms below [21–23]. This shading inhibits the growth of other autotrophic organisms required for wastewater treatment and the removal of coliform bacteria and viruses by natural radiation. Cyanobacterial blooms may also alter the BOD, either by inhibiting natural wastewater treatment or by increasing the BOD when cells decay. This in turn decreases dissolved oxygen concentrations, which can have dire effects on aquatic species when effluent is discharged to the natural environment [10, 24]. Cyanotoxins are harmful to the aquatic biota involved in WSP treatment, including other phytoplankton, zooplankton and protozoa (reviewed in [220]), which can negatively impact treatment processes. These changes in pond ecology caused by cyanobacteria have the potential to cause a shift away from beneficial wastewater treatment organisms, thus inhibiting treatment. This decreases the removal of wastewater pollutants such as coliform bacteria, nutrients and BOD, increasing the risk that insufficiently treated wastewater effluent will be discharged to reuse or the environment. In highly eutrophic, shallow WSPs, prevention of blooms through these measures may not be practical or possible. Other preventative measures may also be of limited use in the WSP environment. The establishment of macrophyte communities has been shown to lower cyanobacterial concentrations, but is impractical in WSPs as macrophyte communities may impact upon wastewater treatment processes and provide breeding habitats for insects that carry vector-borne disease. Predation by the addi-

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tion of zooplankton and fish, and biomanipulation by the addition of decaying organic matter and allelopathic chemicals, have been successful in some reservoirs, but may alter WSP ecology or significantly increase sludge production, further decreasing WWTP efficiency. Ultrasonication to prevent the formation of large cyanobacterial blooms has shown some promise in reduced-scale experiments, but is yet to be trialed on scales large enough to infer its preventative efficiency at the full-scale. Ultraviolet radiation may prevent cyanobacteria in some situations, but where cyanobacteria are already exposed to UVR they have likely developed defense mechanisms, reducing the suitability of UVR as a preventative technique. Waste stabilization ponds are eutrophic, shallow systems which experience high levels of natural irradiance, so it is unlikely that cyanobacterial blooms can always be prevented. Consideration of the cyanobacterial and cyanotoxin removal methods studied in the literature (Tab. 10.3) suggests that hydrogen peroxide (H2 O2 ) may be suitable for reducing cyanobacterial and cyanotoxin concentrations, and may be more successful when coupled with other mitigation techniques. Hydrogen peroxide degrades within hours of addition [221–224], is not considered to be carcinogenic [225] and is unlikely to impact significantly upon aquatic biota at the concentrations required for cyanobacterial removal [177, 226, 227]. Hydrogen peroxide should not pose a risk to ecosystems or humans if treated effluent is discharged to the environment or for reuse. Hydrogen peroxide has been used occasionally for cyanobacterial management in WSPs, although there has been minimal scientific investigation into the removal dynamics of cyanobacteria and cyanotoxins by this method. The addition of H2 O2 alone has often been considered inadequate for cyanotoxin treatment [140, 165, 170, 174, 228]. However, there are many physical, chemical and biological properties of WSPs that differ from laboratory studies, and the presence of such factors may improve the potential for cyanobacterial and cyanotoxin treatment by H2 O2 under WSP conditions.

10.7 Treatment of cyanobacteria and cyanotoxins with hydrogen peroxide Cyanobacterial and cyanotoxin removal by H2 O2 proceeds via the generation of hydroxyl (•OH) and hydroperoxyl (•OOH) radicals and is illustrated in Fig. 10.8. Hydroxyl and hydroperoxyl radicals are produced naturally from H2 O2 through interaction with chemical catalysts (e.g. iron) and UVR. These radicals damage cells via multiple pathways, including membrane disruption, mutagenesis, bleaching of pigments, oxidation of photosystem II, reduction of carbon dioxide fixation and the division of peptides [164, 166, 179, 229–233]. These mechanisms lead to oxidative stress within cells, similar to the effects caused by photo-inhibition under high natural radiation doses [234]. Most phytoplankton are able to repair systems damaged by photo-

10 Control and management of Harmful Algal Blooms

1. Fe2+ / Fe3+

2.

3.

H2O2

Cyanobacterial cell: Oxidative stress •OH production

•OH

327

UVR + PAR

•OH •OOH

UVR

|

Cyanobacterial cell: Oxidative stress •OH production

•OOH

4.

5.

6.

COOH

UVR

Cyanobacterial cell: Cellular lysis Cyanotoxins released

Natural pigments O N

N

NH COOH

O

O NH O NH2 HN

O

O N

N

NH

NH2

CH2 O

H N

O NH2

NH CH2

HN

O H N

H N O COOH O

HN

NH CH2 O

H N

H N

HN

O COOH O

N H

OH

HN

H N O COOH

O

O

O

O NH

COOH

•OH OH

OH

O N

N

HN O

N H

N H

Fig. 10.8: Method of action of hydrogen peroxide (H2 O2 ) for cyanobacterial and cyanotoxin removal. 1. H2 O2 is added to the water column. Chemical catalysts (e.g. iron) and ultraviolet radiation (UVR) react with H2 O2 to produce hydroxyl (•OH) and hydroperoxyl (•OOH) radicals. 2. •OH and •OOH radicals attack cyanobacterial cells, inducing oxidative stress and an increased cellular production of •OH. 3. Oxidative stress may lead to death if cellular processes are unable to repair the damage to core systems. Damaged cells become more susceptible to damage by UVR and photosynthetically active radiation (PAR). 4. Cell death and lysis occurs. This releases cyanotoxins (here microcystin-LR) into the water column. 5. •OH and •OOH radicals attack the conjugated bonds of cyanotoxins, which can lead to oxidation and ultimately cleavage. UVR interacts with natural pigments in the water column, which can lead to isomerization of the cyanotoxin molecule. 6. The toxicity of the cyanotoxin is destroyed.

inhibition and oxidative stress within hours [229, 232–235]. Where these mechanisms are insufficient, permanent damage occurs, and it is likely that core photosynthetic activities have been lost, resulting in cyanobacterial death. The ensuing cell lysis releases cyanotoxins to the water column. Hydroxyl and hydroperoxyl radicals destroy the toxicity of cyanotoxins by targeting their conjugated diene structure, forming dihydroxylated products and inducing cleavage of the molecule [236]. Other natural processes such as thermal decomposition, isomerization, adsorption and biodegradation also destroy the toxicity of dissolved cyanotoxins, and studies have determined that oxidation of cyanotoxins results in non-toxic by-products [180, 188, 237]. Hydrogen peroxide may thus be effective for treating both cyanobacteria and cyanotoxins in WSPs.

328 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

Most studies into the use of H2 O2 have been conducted within the laboratory, often on batch cultures of single cyanobacterial species [e.g. 168, 140, 164, 171, 166, 237] or purified cyanotoxin compounds in distilled water [e.g. 172, 173, 165, 170, 227, 174, 175]. The results of such investigations are not necessarily indicative of the dynamics of cyanobacteria and cyanotoxins following H2 O2 addition to WSPs. The effects of H2 O2 on mixed phytoplankton assemblages under environmental conditions may be altered by a number of physical, chemical and biological variables. Natural irradiance increases the concentration of •OH and •OOH in solution [168] and further damages cells already inhibited by oxidation [239]. Mixing and stratification alter the position of phytoplankton and algicides within the water column, directly affecting the treatment of cells [240, 241]. The presence of other biota and compounds that may react preferentially with •OH and •OOH may reduce the removal of cyanobacteria and cyanotoxins by H2 O2 , and the dynamics of H2 O2 as an algicide may differ between prokaryotic cyanobacteria and eukaryotic phytoplankton species present in the assemblage [166, 168, 177, 242]. Although H2 O2 does induce cell death in various phytoplankton, the decay of cyanobacterial cells appears to occur more rapidly than eukaryotic phytoplankton, suggesting that H2 O2 may be a selective algicide when applied under environmental conditions [177]. Such environmental variables must be investigated to determine the applicability of H2 O2 as an algicide. Many studies into the effectiveness of H2 O2 for removing cyanobacteria and cyanotoxins have been conducted on short timescales and at low temporal resolution (e.g. [140, 165, 167, 168, 170, 172–175, 228, 243]). In these investigations, particularly where H2 O2 has not been coupled with other physical or chemical mitigation techniques, the removal of cyanotoxins by H2 O2 has been considered negligible [140, 165, 170, 174, 228]. Short measurement periods have been considered sufficient given the rapid decay of H2 O2 and the requirement for fast removal of cyanobacteria and cyanotoxins in rapid flow-through systems such as drinking water treatment. However, in systems such as WSPs, immediate cyanobacterial cell death and cyanotoxin removal is not required. Where H2 O2 induces oxidative stress, cells may die within a timescale longer than that of H2 O2 decay, releasing cyanotoxins to the dissolved state. Natural mechanisms may then degrade dissolved cyanotoxins given sufficient retention time, and it is thus important to monitor the effectiveness of H2 O2 addition on timescales that allow for the induction of cellular stress followed by death, as well as natural degradation of cyanotoxins. In WSPs, water can be retained for several days following algicidal treatment, and the degradation of cyanobacteria and cyanotoxins over a longer timeframe than those traditionally investigated may be suitable. Cyanobacterial treatment methods that result in the release of cyanotoxins from cells (e.g. [110, 244–246]) have generally been considered unfavorable for water management. Most past studies into the use of H2 O2 have been conducted within the laboratory, so the degradation of cyanotoxins by natural mechanisms following H2 O2 addition has not been thoroughly investigated. Cyanotoxins may be degraded more

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rapidly from the dissolved state than whilst cell bound [247], suggesting that cyanobacterial mitigation techniques that induce cell lysis may be suitable where dissolved cyanotoxins can be degraded naturally. The methods of detoxication vary between water bodies, but studies indicate that thermal, photolytic, adsorptive and biodegradation of cyanotoxins occur within the environment. Thermal decomposition of cyanotoxins is generally most effective at acidic pH values [248], which are rarely encountered in WSPs. Photolytic degradation can occur rapidly in systems containing high concentrations of pigments, but it is often not a significant removal mechanism in the short-term [191, 249, 250]. Adsorption onto natural matter may be unreliable for ensuring cyanotoxin degradation, particularly where cyanotoxins are suspended and not undergoing filtration [114–117, 119, 122]. Biodegradation has shown considerable success in degrading cyanotoxins (reviewed in [251]) (Tab. 10.3). Complete cyanotoxin removal has been observed to occur within two to three weeks under most environmental conditions. Cyanotoxins may act as substrates for certain bacteria, so the increase in bacterial populations may increase the rate of cyanotoxin removal. This suggests that higher concentrations of cyanotoxins may be biodegraded more rapidly than lower concentrations [252]. If natural biodegradation processes can be relied upon for the removal of cyanotoxins from WSPs, the expensive toxin removal techniques often utilized in drinking water treatment will be unnecessary. The use of H2 O2 for the removal of cyanobacteria and cyanotoxins from reservoirs introduces a larger scale than that investigated through laboratory experiments. Although controlled microcosm experiments are important when initially investigating the use of cyanobacterial and cyanotoxin removal methods, it is difficult to replicate environmental phenomena and heterogeneity at smaller scales. Such characteristics may significantly impact the use of H2 O2 as a mitigation technique. The position of cyanobacterial cells and cyanotoxins within the water column is an example of a larger-scale phenomenon that may impact H2 O2 use. Cells and cyanotoxins may be influenced by buoyancy regulation or stratification [253, 254], which themselves depend upon temperature, radiation and wind conditions [255], phenomena not often included in laboratory-scale studies. In order to test the true management potential of H2 O2 as an algicide, scaled field trials are required to infer the differences in cyanobacteria and cyanotoxin dynamics following full-scale application [256–258]. Hydrogen peroxide has shown promise in reducing cyanobacteria and cyanotoxin concentrations in multiple studies. However, there has been limited investigation into the use of H2 O2 for removing cyanobacteria and cyanotoxins from natural phytoplankton assemblages, particularly under field conditions and at the reservoir-scale, and no previous work has investigated the dynamics of cyanobacteria and cyanotoxin removal using H2 O2 in WSPs. Should H2 O2 be determined to be a suitable method for the removal of cyanobacteria and cyanotoxins from WSPs, a framework for the management of cyanobacterial blooms within WWTPs using H2 O2 may be developed for use by water utilities.

Bacterial source

River

River

River

Lake

Lake

Lake

Reservoir water and bed sediment

River and loch

Study

[259]

[259]

[260]

[261]

[262]

[262]

[263]

[264]

Extract from culture

Stock solution

Lysed natural algal material

Lysed natural algal material

Extract from natural bloom

Extract from natural bloom

Extract from natural bloom

Extract from natural bloom

Toxin source

River and loch water (L)

Reservoir water and bed sediment (L)

Lake water (F)

lake water (L)

culture medium (L)

river water (L)

tap water with biofilm (L)

river water with biofilm (L)

Experimental environment

Tab. 10.3: Studies into the biodegradation of cyanotoxins.

—

AD

ND

D

—

—

AD (0/12 μE m−2 s−1 )

AD (0/12 μE m−2 s−1 )

Light regime

1000–5000 μg MC-LR/-RR/-LW/ -LF/NOD l−1

10 μg MC-LR l−1

2–54 μg MC-LR eq. l−1

10–136 μg MC-LR eq. l−1

0.7 μg MC-LR l−1 / 1.7 μg MC-RR l−1

50 μg MC-LR l−1

160 μg MC-LR/-YR l−1

160 μg MC-LR/-YR l−1

Initial cyanotoxin concentration

29

17–21

15.5–21.5

20

30

—

22

22

Temp. (°C)

100 % reduction in 7–19 days

100 % reduction in 6–7 days

Decreased to < 1 μg/l after 1–4 days, reduced to detection levels after 8 days

Decreased to < 1 μg/l after 7 days, 100 % reduction in 21 days

100 % reduction in 24 hours at pH 7, less degradation at other pH values

90 % reduction in 2 days, 100 % reduction in 12 days

Half-life of 17–84 hours

Half-life of 20–23 hours

Results

330 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

Bacterial source

Lake (during cyanobacterial bloom)

Lake (no cyanobacterial bloom present)

Biological sand filter

Tertiary treated effluent

Activated sludge treated effluent

Lake sludge

Lake

Lake

Lake bloom treated with copper sulfate

Study

[265]

[265]

[163]

[252]

[252]

[266]

[267]

[268]

[244]

Tab. 10.3 (continued)

Lake water (L)

—

Lake bloom treated with copper sulfate

Lake (naturally occurring)

Extract from natural bloom

Extract from culture

Stock solution

Stock solution

0.14–8.93 μg MC-LR/-RR l−1 1300–1800 μg MC-LR eq. l−1

ND

Lake bloom treated with copper sulfate (F)

6000 μg MC-RR/NOD l−1

Variable

Variable

30

30

10–22

20 μg MC-LR l−1 ≈ 2–4 μg MC-LR l−1 / ≈ 4 μg MC-RR l−1

22

6–20 μg MC-LR l−1

22–30

20–25

20 μg NOD l−1

3–25 μg MC-LR/-LA l−1

8–25

Temp. (°C)

20 μg NOD l−1

Initial cyanotoxin concentration

ND

D

—

—

—

—

AD

AD

Light regime

Mesocosms within lake (F)

Culture medium (L)

Culture medium (L)

Tertiary treated effluent (L)

Tertiary treated effluent (L)

Reservoir water (L)

Lake water (L)

—

Extract from natural bloom

Experimental environment

Toxin source

9 day lag phase before degradation began, 94 % degradation after 12 days

Toxin removal varied over several months

100 % reduction in 4 days

100 % reduction in 4–17 hours

100 % reduction in 7–22 days

100 % reduction in 3–4 days

100 % reduction in 2–15 days

100 % reduction in 7 days

100 % reduction in 2–15 days

Results

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331

Bacterial source

Culture

Culture

Culture

Lake

Lake

Final effluent from activated sludge WWTP

Lake

Estuarine water

Study

[269]

[269]

[269]

[110]

[270]

[271]

[272]

[273]

Tab. 10.3 (continued)

Stock solution or lysed culture material

Stock solution

25

14–15

10 000 μg MC-LR/LF/-LW/-LY/-RR/ NOD l−1 ≈ 1000 μg MC-LR/ [D-Leu1 ]MC-LR l−1

D

—

Estuarine water (L)

25

Variable

20

Filtered and sterilised lake water (L)

0.06–3.2 μg MC-LR l−1

≈ 1 μg MC-LR l−1

20

20

1000 μg MC-LR l−1 1000–16 000 μg MC-LR l−1

20

Temp. (°C)

1000 μg MC-LR l−1

Initial cyanotoxin concentration

210–1620 μg MC-LR l−1

Sewage effluent (L)

Extract from culture

D/ND

AD

—

—

—

Light regime

AD

Mesocosms within lake (F)

Lake water treated with copper sulfate (L)

River water (L)

Lake water (L)

Drain and dam water (L)

Experimental environment

Lake

Lake

Extract from natural bloom

Extract from natural bloom

Extract from natural bloom

Toxin source

100 % reduction in 10–20 days

Half-life of 5–> 10 days

Undetected by day 13–27

90 % reduction in 15–30 days

MC released to dissolved phase over 4 days Half-life of 3 days following maximum release

Degradation rates increased with initial MC concentration > 95 % reduction in 23 days

No significant degradation in 12 days

3–6 day lag phase before degradation began, >95 % reduction in 10 days

Results

332 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

Bacterial source

Culture

Culture

Culture

Lake

Lake

Lake

Cultures

Lake

Study

[274]

[275]

[276]

[277]

[277]

[278]

[279]

[280]

Tab. 10.3 (continued)

Stock solution

—

—

Extract from natural bloom

Extract from natural bloom

Extract from culture

Extract from culture

Extract from culture

Toxin source

Culture medium (L)

Culture medium (L)

Culture medium (L)

Culture medium (L)

Culture medium (L)

Culture medium (L)

Culture medium (L)

Culture medium (L)

Experimental environment

—

—

D

D

—

—

—

—

Light regime

50 000 μg MC-LR l−1

100 μg MC-LR l−1

1000 μg MC-LR/-RR/-YR l−1

3–37 μg MC-LR/-RR/-YR l−1

20 μg MC-LR/-RR/-YR l−1

100–4000 μg MC-LR/-RR l−1

Various concentrations of CYN and MC-LR/-RR/ -LW/-LY/-LF

1/10/100 μg MC-LR l−1

Initial cyanotoxin concentration

25

22/37

30

5–30

27

95.5 % reduction in 21 days

Varied between bacteria, greatest reduction 80 % in 25 hours

95–98 % reduction in 2.5 hours

10–30 °C, 100 % reduction in 6 days 5 °C, 60 % reduction in 7 days

100 % reduction in 6 days

99–100 % reduction in 5–9 days

Varied. Greatest reductions 60.3 % MC-LR, 62.8 % MC-RR, 77.4 % MC-LF, 31.6 % CYN

37

—

Varied significantly between bacterial strains. Greatest reduction 60.1 % in 24 hours

Results

4/22/37

Temp. (°C)

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Lake water and sediment

Reservoir

Lake sediment

Culture

[281]

[282]

[283]

[284]

Culture

Stock solution

—

Stock solution and extract from natural bloom

Toxin source

—

—

114 μg MC-LR l−1

42 300 μg MC-RR l−1

200 μg MC-LR/-RR l−1

200 μg MC-LR/-RR l−1

Initial cyanotoxin concentration

—

30

—

27

Temp. (°C)

Up to 82.7 % removal in 40 hours

100 % degradation in 10–36 hours

100 % degradation in 36 hours

< 1 day for total degradation

Results

—: indicates that details of the property were not stated.

Concentration: MC-LR: microcystin-LR, MC-YR: microcystin YR, MC-RR: microcystin-RR, MC-LW: microcystin-LW, MC-LF: microcystin-LF, MC-LA: microcystin-LA, [D-Leu1 ]MC-LR: [D-Leu1 ] microcystin-LR, NOD: nodularin, CYN: cylindrospermopsin, MC-LR eq.: microcystin-LR equivalents.

AC: artificial constant.

Culture medium (L)

Culture medium (L)

—

D

Culture medium (L) Culture medium (L)

Light regime

Experimental environment

Light regime: D: dark, AD: artificial diurnal, ND: natural diurnal,

Experimental environment: L: laboratory, F: field.

Bacterial source

Study

Tab. 10.3 (continued)

334 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

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10.8 New techniques for the control and characterization of cyanobacterial blooms 10.8.1 Allelopathic control of cyanobacteria As discussed previously, there are many control methods for blooms; however, in some situations only a few of them are applicable due to other factors, including secondary pollution, high cost or no-target ecosystem effects. Consequently, there is a need for anti-algal agents that are more specific, environmentally-friendly and cost-effective. Allelopathy is the direct or indirect effect of plants (including microorganisms) on others through the production of chemicals. This technique could be utilized for the development of an anti-algal agent for the control of harmful algal blooms; the allelopathic activity of barley straw to many kinds of algae including Microcystis has already been documented in the field and laboratory [58, 285–287]. Barley, Hordeum vulgare L., is one of the earliest cultivated crops in the world and can be divided into two distinct groups: occidental and oriental. Most studies on cyanobacteria control using barley were undertaken in Europe and America with the cultivated occidental barley as an anti-algal agent. Oriental barley, i.e. Tibetan hulless barley (Hordeum vulgare L. var. nudum) originated from the Qinghai–Tibetan Plateau and is regarded as the progenitor of cultivated barley [288]. One of the key interests in allelopathy in China is to assess the acute, mid and long-term effects of Tibetan hulless barley straw extract on the growth, physiology and morphology of Microcystis aeruginosa at a single cell level [54, 63]. Recent data shows that a dosage of 2.0 g (dry weight) l−1 of Tibetan hulless barley straw reduces the in vivo chlorophyll-a (chl-a) fluorescence of M. aeruginosa cells, resulting in a significant decline of the cell density (Fig. 10.9). These studies show promise for future use of Tibetan hulless barley straw algicidic agent for M. aeruginosa [54, 63, 64].

336 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

0

0.5 4

2 g L–1 barley

120

1 8

Cell density (of % control)

100

80

60

40

20

0 1d

5d Days

15d

Fig. 10.9: Inhibition effect on growth of Microcystis aeruginosa cultures after a 1-, 5-, and 15-d Tibetan hulless barley straw exposure. All error bars correspond to the standard deviation. Reproduced from [54].

10.8.2 Optimization of the FDA-PI method using flow cytometry to measure metabolic activity of cyanobacteria The control of M. aeruginosa blooms has been the focus of many studies because of the widespread occurrence and the spectacular nature of the bloom events (Fig. 10.10), where this particular species can completely dominate the phytoplankton community [289, 290].

Fig. 10.10: Spectacular M. aeruginosa algal bloom in Lake Taihu, China. (Photos: Hohai University, Nanjing, China).

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Advanced methods for the detection and assessment of the physiological status of this particular species have been developed to ensure a reliable diagnosis of the algal blooms. One such method is the exploration of the inhibition of enzyme activity, which is now a widely accepted method to determine acute and sub-lethal endpoints for bioassays, and to assess the integrity of the cell membrane during the blooms [291]. The release of large amounts of endotoxins can constitute a major environmental and human health hazard. Therefore, it is important to develop reliable diagnostic tools to monitor the blooms events, for a better risk assessment [292, 293]. Many kinds of fluoresceins had been used to detect the enzyme activities and cell membrane integrity. Among them, fluorescein diacetate (FDA) and propidium iodide (PI) were most frequently used. FDA is a non-polar, hydrophobic, non-fluorescent esterified compound; it readily permeates the cell membrane, and is hydrolyzed by nonspecific esterases producing a fluorescein [294]. The mean fluorescence intensity per cell (MFI) of FDA-converted fluorescein was used to estimate the the enzyme activity (i.e. hydrolysis rate of esterase) in algae. PI is a fully cell-membrane impermeable fluorescent dye which has been used to indicate dead cells for a wide range of microorganisms [295, 296]; it can only combine with DNA in dead cells or cells with damaged membranes [295, 297]. In contrast to PI, the use of FDA was first reported for detecting the viability of marine phytoplankton after the exposure of environmental contaminants using a fluorescence microscope [298]. Cells with an intact cell membrane are stained bright green by FDA; in contrast, cells with a broken cell membrane are stained bright orange with PI [299, 300]. Furthermore, FDA also indicates the presence of active esterase [294, 300]. Cells with an intact cell membrane and inactive esterase do not stain with FDA or PI. Afterwards, the efficiency of PDA/PI detection was greatly improved by the detection of the fluorescence of individual cells using flow cytometry [301]. More recently, Franklin et al. [291] developed a rapid enzyme inhibition bioassay based on PDA/PI for marine and freshwater microalgae with the use of flow cytometry, but no evidence of M. aeruginosa was detected in this study. Moreover, FDA has been used to evaluate the esterase activity of M. aeruginosa, but dosages used ranged from 1.6 to 16 mg l−1 and the incubation time differed from 8 min to 2 h [302, 303]. Recently, a new procedure based on an optimized FDA/PI condition has been developed for short-term bioassays [54, 304]. This new procedure takes working conditions such as pH and impure cultures into consideration, could avoid algal cell damages in sample preparation and separate algal cells from non-algal particles by fluorescence triggering. This procedure has been used to assess the toxicity of copper on M. aeruginosa in a short-term exposure (36 h). As copper concentrations increased, the esterase activity was found to decrease in a concentration-dependent manner and the membrane fragments increased (Fig. 10.11). Moreover, esterase activity was a good indicator of copper toxicity in M. aeruginosa. The EC50 value based on MFI was 101.5–146.2 μg/l (95 % confidence intervals) [304]. Therefore, this new procedure has the potential to be sub-lethal endpoint detection, and could be used for the selec-

100

101

102 FL3–H

103

104

100

101

102 FL3–H

103

104

SSC–H 100

101

SSC–H

103

100

104 (C)

101

102 FL3–H

103

104

0

0

102 FL3–H

200 400 600 800 1000

(B)

SSC–H 200 400 600 800 1000

(A)

100 (D)

0

0

0

SSC–H 200 400 600 800 1000

SSC–H 200 400 600 800 1000

(CK)

200 400 600 800 1000

0

SSC–H 200 400 600 800 1000

338 | Dani J. Barrington, Xi Xiao, Liah X. Coggins, and Anas Ghadouani

101

102 FL3–H

103

104

100 (E)

101

102 FL3–H

103

104

Fig. 10.11: Flow cytometry images of M. aeruginosa cells after 36 h copper exposure by Side Scatter (SSC) and FL3 detectors. (CK: blank control; A: 25 μg (Cu)/l; B: 40 μg (Cu)/l; C: 63 μg (Cu)/l; D: 100 μg (Cu)/l; E: 158 μg (Cu)/l).

tion of M. aeruginosa control methods or investigation of the M. aeruginosa activity inhibition mechanism as a rapid and cost-effective bioassay.

10.9 New perspectives and future directions In recent years, the study of waste stabilization ponds (WSPs) as ecological systems has revealed new considerations that are likely to influence how we see a wide range of aquatic systems with cyanobacteria including a lakes, rivers and reservoirs. In particular, it was determined that the interplay between the hydrodynamics and ecology within these systems can explain the occurrence, magnitude and frequency of cyanobacterial blooms (Fig. 10.12). Sludge accumulation can impact performance by reducing pond effective volume and changing the shape of the bottom surface, thus altering pond hydraulics [205,

10 Control and management of Harmful Algal Blooms

Chemical treatment

Environmental Risk

339

Green algae

–

(System state sensu Scheffer et al. 2001)

|

De-sludging

System Reset

Desirable stable state (DSS) Critical operation point

Unstable state (Regime shift region)

Critical threshold (Break Point)

Undesirable stable state (USS) Toxic cyanobacteria

+ +

Hydraulic performance

–

Fig. 10.12: Conceptual framework inspired by Scheffer’s theory of catastrophic shift in ecosystems applied to a situation of intense eutrophication of WSP systems. The shaded area (lines) represents the desirable operational state of the system. Below the critical operation point small changes in hydraulic performance can lead to catastrophic shifts, driving the system into an undesirable stable state. A similar situation could be expected in hypereutrophic systems where intense cyanobacterial blooms occur. Reproduced from: Ghadouani and Coggins [218].

305, 306]. While periodic sludge removal is required, it is rarely considered integral to pond design [205], and the long-term sustainability of WSP systems is dependent on the safe management of sludge [206]. Previous studies have shown that distribution of sludge in ponds can be very uneven [205, 206, 217] (Fig. 10.13) and that different climatic regions have an effect on sludge accumulation rates [206]. Despite the number of WSPs worldwide (e.g. in regional Western Australia there are 84 wastewater treatment plants using 302 WSPs for treatment), there is still little information available on sludge distribution, sludge characteristics, accumulation rates, and their effect on wastewater treatment efficiency. Besides sludge accumulation, factors that influence pond hydraulic performance are mainly related to shape, flow, inlet/outlet configuration, wind and temperature [307–310]. Optimal flow within treatment systems is described as flow with a uniform velocity profile, and it is recommended that ponds be designed to adhere to plug flow [307]; this flow regime provides mean maximum residence time. However, in reality, water in ponds does not move homogeneously, but with eddies and recirculation [307], and the actual mean residence time is always less than the nominal residence time [308] Microbial and phytoplankton communities are essential for the functionality of WSPs, and community health is important for overall treatment efficiency. Microbial processes in ponds, such as algal growth, aerobic and anaerobic heterotrophic

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metabolism, nitrification and denitrification, work in conjunction with physical processes and exposure to sunlight to remove pathogens, nitrogen and organic contaminants [207]. The highly complex microbial communities present in wastewater treatment systems are not well understood, despite their importance in the treatment process [311]. Recent studies have indicated that there is a link between microbial diversity, community structure and treatment efficiency [312, 313], and that total bacterial cell counts [314] and phytoplankton presence [315] can be used as descriptive parameters for treatment processes and performance.

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Joaquín Espinosa, Sara Silva-Salvado, and Óscar García-Martín

11 Global climate change profile and its possible effects on the reproductive cycle, sex expression and sex change of shellfish as marine toxins vectors 11.1 Introduction Marine bivalve mollusks (shellfish) are organisms with an ubiquitous distribution in the world oceans; they feed by filtering phytoplankton and dissolved organic matter in seawater. These organisms are high in protein, glycogen and lipid, as well as in hydro- and fat-soluble vitamins, carotenoids and essential minerals. Because of their feeding, mostly microalgae, these organisms are also high in polyunsaturated fatty acids (PUFAs). For all these reasons, bivalve mollusks have high nutritional value and are essential in the diet of populations in many coastal areas of the world, mainly in developing countries that are the ones that commercialize larger volumes of these animals. Because bivalve mollusks are filter feeders, the organisms accumulate molecules in their body that originate from the organisms they feed, especially various species of phytoplankton that produce, depending on environmental conditions, molecules (toxins) that are toxic for humans. When toxic algal blooms reach significant volumes, they are called episode HABs; and in recent years these phenomena have been more frequent. Global climate change is an incontestable fact today. Natural and anthropogenic causes are provoking major changes in physical, chemical and biological parameters in all ecosystems worldwide, including oceans. Temperature, salinity, CO2 concentration, pH, oxygen concentration, light, nutrients and primary production are characteristics of the marine environment that are being affected, significantly and in a very short time, by global climate change. Biological characteristics of marine organisms, including bivalve mollusks, are also affected by these changes; reproductive cycles, sex expression and sex change may be some of these features. In the present study, the profile of global climate change is exposed, and its effects on adaptability, reproductive cycle, sex expression and sex change of bivalve mollusks as marine toxins vectors are analyzed.

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11.2 Shellfish as marine toxins vectors 11.2.1 General considerations Shellfish are organisms with a high content of protein, glycogen and lipids whose intake also provides both hydro- and fat-soluble vitamins (particularly A and D), essential minerals and carotenoids. Shellfish feed mainly on marine microalgae [1]. Algae, and microalgae particularly, are important for its high content in long chain polyunsaturated fatty acids (PUFA) because they have biosynthetic capacity for these molecules; the importance of shellfish in a human diet relies especially upon this high content of PUFAs [1, 2]. Some types of shellfish, such as bivalve mollusks, filter large volumes of water daily, feeding on both organisms and the dissolved organic matter it contains; as a result, they accumulate and concentrate dissolved molecules and organisms present in water, such as pathogenic bacteria and phycotoxins [1, 3]. It has been determined that various human diseases associated with shellfish consumption seem to come from toxins produced by marine microalgae. When microalgae populations multiply rapidly, this can cause visible microalgae blooms because they may have certain colors; these blooms receive the traditional name of “red tides” [1]. These dense concentrations of microalgae are not always visible because either they are colorless or grow in deep water far from the sea surface, and in the latter case they can also not be seen. Because of their toxicity, these microalgae blooms are also known as “harmful algal blooms” (HABs) [1]. Such events can have negative consequences for environmental conditions, as they often lead to decreased oxygen concentration in the water column, and if ingested by fish can cause gill damage [1, 4]. Furthermore, the toxins produced by such algae can result in mass mortalities of fish, birds and marine mammals, as well as diseases in humans through the consumption of seafood products [1]. It has been estimated that of the approximately 4000 known species of marine phytoplankton, only 60–80 are potentially toxic and capable of producing HABs [1, 5]. The problem arises with species of organisms producing toxins (HAB species) able to produce toxic compounds for other species including humans [6]. The route through which these HAB species cause disease in humans is usually indirect and involves the consumption of fish and shellfish that acquired the toxins by ingesting toxic microalgae. Vector organisms of these toxins that are not shellfish include fish (e.g. flatfish), which can be poisoned with a reduced number of toxic compounds that develop high toxicity in fish such as ciguatera toxin organisms (ciguatera fish poisoning, CFP). It particularly affects fish from tropical waters [6]. More common are other diseases caused by HAB species that occur when humans ingest toxins produced by microalgae when they eat shellfish [6, 7]. These diseases can result in different symptoms, from gastrointestinal distress to more serious effects

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caused by toxins with neurological action that can endanger the lives of people who have ingested them. Such diseases have been grouped into five main categories according to the syndrome caused to humans, and they are [6]: (1) paralytic shellfish poisoning (PSP), caused by the saxitoxin group; (2) diarrheic shellfish poisoning (DSP), caused by the okadaic acid group; (3) amnesic shellfish poisoning (ASP), caused by the domoic acid group; (4) neurotoxic shellfish poisoning (NSP), caused by the brevetoxin group; and (5) azaspiracid shellfish poisoning (AZP), caused by the azaspiracid group. At present, there is no procedure to remove toxins from shellfish except the natural removal of them from these organisms, through their own metabolic processes in the same marine zones where shellfish grow. The duration of this natural toxins removal ranges from several weeks to several months, depending on the amount of toxin that is in the water and the environmental conditions thereof [6]. Trade in bivalve mollusks is a crucial economic activity in many coastal areas of the world, particularly in developing countries that remain the largest exporters of shellfish, fish and fishery products, including bivalve mollusks [8, 9]. On average, global production of bivalve mollusks has increased from 5.3 million tons in 1990 to 14.6 million tons in 2010 [8]. This growth is due mainly to the strong increase produced by aquaculture, which has grown from 3.3 million tons in 1990, which represented 62 % of total world production of bivalve mollusks, to 12.9 million tons in 2010, representing 88 % of total world production of bivalves [8]. The economic value of aquaculture products in this period has increased proportionately from 3500 to 13 000 million USD in 2010. However, in this period catches have declined slightly from 2.0 to 1.7 million tons [8, 10]. The expansion in the aquaculture of bivalve mollusks has been particularly notable in Asia, where bivalve production has almost doubled in the last 20 years, from 6.1 million tons in 1990 to 11.6 million tons in 2010. China is by far the largest worldwide aquaculture producer of bivalve mollusks, with 10.3 million tons produced in 2010; Japan, South Korea and Thailand are also major producers. The production of bivalve aquaculture in Europe has remained fairly stable during this period, although there has been a slight decrease from 710 000 to 633 000 tons [8, 10]. Clams, cockles and ark shells are the most important bivalve groups, with 5.5 million tons produced in 2010, followed by oysters with 4.6 million tons in the same period. Scallops account for more than 2.5 million tons, and mussels for 1.9 million tons produced in 2010 [8, 10]. Toxic episodes caused by the presence of marine biotoxins described above have a direct impact on the production and commercialization of marine bivalves worldwide, not only on shellfish harvesters and the aquaculture industry, but also on coastal economies and on human health [8].

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11.2.2 Global increase in HABs The apparent global increase in the occurrence of algal toxins in shellfish and their causes has been considered by different authors in the last 20 to 25 years [1, 11–16]. Although the proliferation of toxic algae as a natural phenomenon is known to have happened long ago [1, 11], several authors that have studied this phenomenon generally agree that there has been an increase in the impacts of these events on the economies and public health from different communities in various parts of the world in the last three decades. The causes of the increase of these impacts can be different and, almost certainly, dependent on the types of toxins and microalgae that produce them. In 1993 Gustaaf Hallegraeff [11] identified four main causes for the global increase in HABs: (1) the increasing of scientific interest in studying and monitoring species of microalgae producing toxic substances, mainly because of damage caused to human health and aquaculture farms; (2) the increased use of different coastal regions worldwide for the installation of aquaculture farms (mollusks, crustaceans, fish and algae) because overfishing resulted in decreased production of these marine species; (3) the rise in phytoplankton blooms as producers of toxic substances caused by the so-called “cultural eutrophication” of the marine environment and/or uncommon climatic conditions, highlighting the changes in the concentrations of basic nutrients in the marine environment (nitrogen, phosphorus, silicon and oxygen) caused by domestic, industrial and agricultural effluents, as well as by the influence of the recurring – but infrequent – climatic events such as the phenomenon of El Niño Southern Oscillation (ENSO); and (4) the transport of dinoflagellate cysts caused either by the emptying of ballast water from ships or moving shellfish from one place to another. In 2000, van Dolah [12], and Masó and Garcés in 2006 [13], also reviewed the environmental conditions that could favor the rise of HAB episodes and agreed with Hallegraeff [11] that transport – caused either by the emptying of ballast water from ships or moving shellfish from one place to another – is an important factor, but Masó and Garcés [13] introduced a variation also attributing this capacity for transport and distribution of microalgae in marine environment to the floating plastic waste [13, 17]. These authors also agreed with Hallegraeff on other factors: the anthropogenic eutrophication of the marine environment and the occurrence of anomalous climatic events such as the phenomenon of El Niño. Masó and Garcés [13] also noted that the decreased biomass of filter-feeder organisms – due to overfishing – or the change of environmental conditions in a particular area could favor the HAB episodes of certain toxic microalgae [18], as in the case of the Mediterranean Sea, where there is an increase in confined water masses because of overexploitation of the coastline [19, 20]. However, the novel factor in both reports is the reference to global climate change [12], with the change component induced by humans [13] as an enhancer factor of HAB episodes. Climate change, although it is mentioned in such reports as a possible cause of HAB episodes, the effect of global climate change is limited to the effect of global warming caused by the increase in the concentration of greenhouse gases; both re-

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ports point out that since there is no broad historic series of data yet, the real impact of global climate change on the proliferation of HAB species cannot be assessed effectively. In 2007 and 2008, the studies making up the 4th IPCC report on global climate change were published [21, 22], removing any doubt about the reality of the changes taking place in the Earth’s climate. The contributions of K. J. James et al. [1] and G. Hallegraeff in 2010 [14], while still maintaining other factors as causative for the increase in proliferative episodes of HAB species, especially the report of K. J. James et al., factors such as anthropogenic activities, transfers of cultured shellfish, eutrophication, increased global marine traffic, increased ability to detect marine toxins, better food control and toxins monitoring programs [1, 12, 13, 23–25], climate change is at the forefront among the possible triggers of the global increase of HAB episodes. James et al. [1] point out that the increase in the concentration of greenhouse gases will result in: (a) lowering the pH of ocean water; (b) increasing the temperature of the surface layer of the oceans; and (c) it will also affect the vertical mixing of waters and upwelling [1, 26]. However, as the report findings manifest, the main causes of the apparent increase in HAB episodes in recent years were the activities of anthropogenic origin that might be called “classic” or “traditional”, i.e. eutrophication of coastal zones [1, 13, 27, 28], marine transport [29] and aquaculture [30, 31]. In this regard, global climate change is referred to as “has also been implicated”. The approach taken by G. Hallegraeff is different in his revision of the various causes affecting toxic algal blooms [14]. Almost all the causes the author considers may be included within the effects that global climate change is having on the world’s oceans. They include: temperature rise, increased surface stratification, alteration of ocean currents, intensification or weakening of nutrients upwelling, stimulation of photosynthesis by elevated CO2 concentrations, reduced calcification because of ocean acidification (the “other” CO2 problem), changes in the supply of terrestrial water to the oceans and variations in the availability of micronutrients. The approach the author gives in this study is original, because it proposes that changes in phytoplankton communities may be used as a sensitive and early warning of disturbances that global climate change is having on marine ecosystems. This aimed to tackle one of the biggest problems facing human societies that, in the author’s opinion, is not currently being prepared to face a significant increase in the proliferation of HAB species and the consequent problems caused by toxins from these algae, particularly in areas of the world with little or no oversight in these issues. The author also suggests that combined results from monitoring the interconnections between toxic algal blooms, changes in phytoplankton communities and the effects of climate change on the oceans over extended periods of time could be a tool of great importance to predict the impact of global climate change on HAB episodes [14]. The last two references quoted in this section [15, 16], review the effects of global climate change on topics different from HAB episodes but related to them, thus completing an overview of the potential effects that global climate change might have on

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the main chain links of primary productivity in oceans and freshwater. The first of these studies [15] analyzes the causes, consequences and control of dangerous proliferation of cyanobacteria in both oceans and freshwater. These organisms, one of the first living beings that appeared on our planet (3.5 bya), were responsible for modifying the reducing composition of the primitive atmosphere of the earth, providing much oxygen to it and thus being responsible for the terrestrial colonization of plants and animals [15, 32]. At present, cyanobacteria have ecophysiological strategies that allow them to exploit the anthropogenic modifications of their aquatic environment (marine and freshwater), which has led them to a relative ecological success manifested in the increasingly frequent and visible proliferations of harmful cyanobacteria, or CyanoHABs [33]. The “harmful” aspects of these proliferations start with the loss of water clarity leading to the elimination of macrophytic organisms and negatively affecting the populations and habitats of invertebrates and fish. Further decomposition of these large numbers of organisms leads to decreased local oxygen concentrations, causing areas of hypoxia or anoxia and therefore the death of other living organisms (invertebrates and fish). In addition, many CyanoHABs produce toxic secondary metabolites (cyanotoxins) that may cause acute poisoning in mammals, including humans [15, 34–36]. Although today one considers water eutrophication to be the main factor in promoting the onset of CyanoHABs [15, 33, 37], various effects of global climate change (global rise in temperature, changes in climatic conditions and precipitation patterns, and salinization) also act as catalysts for its expansion [38–40]. The last study to be considered under this heading was conducted by H. G. Dam on marine zooplankton’s evolutionary adaptation to global climate change [16]. Zooplankton plays a pivotal role in the oceans trophic chain [41, 42], therefore its adaptability to any changes occurring in its natural environment is essential for maintaining the integrity of the whole ecosystem, from its predators (fish and other organisms) to its prey (phytoplankton and others). Furthermore, zooplankton species are organisms that can be used as ideal models to study the responses of animals to changes caused by global climate change, because they have shorter times between generations (weeks to months), thus being able to experience relatively fast evolutionary changes [16, 43]. The author evidences the reality of the adaptive capacity of zooplankton to blooms of toxic algae, hypoxia and some effects – a global rise in temperature of the oceans surface – caused by global climate change. From the perspective of evolutionary ecology, he also reflects on the implications of the adaptive processes as useful tools for understanding and predicting the responses of zooplankton to changes it faces in its environment [16]. In this predictive context, H. G. Dam mentions three recent reviews by George N. Somero [44–46] where the author – in our opinion the world’s most important specialist in comparative physiology and organisms’ adaptive biochemical processes in the marine environment – describes, in his usual clear and elegant way (as appropriately recognized by H. G. Dam), the use of a comparative physiology approach to provide a mechanistic understanding of the adaptation at the organism level while addressing

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the bottlenecks and limitations of this approach for understanding and predicting the responses of all the organisms (biota) to the effects of global climate change. Furthermore, we emphasize that, as stated by G. N. Somero in the above revisions, the study of comparative physiology – as well as the biochemical adaptability and evolution of bivalve mollusks to the different challenges they face today due to the great variability in their environmental conditions – is an indispensable tool for enhancing the above mentioned predictive capability [44–46].

11.2.3 Global climate change 11.2.3.1 Background The studies that make up the 5th Report (AR5) of the Intergovernmental Panel on Climate Change (IPCC) were published in the years 2013 and 2014 [47–67]. The IPCC was created in 1988 by the World Meteorological Organization and the United Nations Environment Programme to provide the governments of different countries studies that supply them with clear views on the current status of knowledge regarding the science of climate change, potential impacts and options for adaptation and mitigation through regular assessments of the latest published information in the scientific, technical and socioeconomic literature worldwide. Three Working Groups – I, II and III – constitute the IPCC. Working Group I provides a comprehensive assessment of the state of knowledge of climate change’s basis in physical science. Working Group II assesses the impacts, adaptation and vulnerability to climate change, and Working Group III studies climate change mitigation. In the 5th IPCC report (AR5) a Synthesis Report [67] is also given that elaborates on the three working groups’ reports, which also include Relevant Special Reports, providing an integrated view of climate change as the final part of the 5th IPCC report (AR5). Although it has already been published, the final version of this Synthesis Report is still pending final processing [67]. Prior to this 5th Report, the IPCC has published four other reports: the IPCC First Assessment Report (FAR, IPCC, 1990) [68], IPCC Second Assessment Report (SAR, IPCC, 1996) [69], IPCC Third Assessment Report (TAR, IPCC, 2001) [70] and the IPCC Fourth Assessment Report (AR4, IPCC, 2007) [71]. All these reports followed the evolution of certain climatic parameters, and two of such variables monitored continuously from the start were the average air temperature on the land surface and the oceans sea level. When comparing the evolution of these two variables through the various IPCC reports, a clear picture of the quality and quantity of global climate change can be attained [49]. Regarding the global air temperature average on the land surface, the first IPCC report (1990) reported that this parameter increased by 0.3–0.6 °C over the last 100 years, with the 1980s being the years with a warmer average. The second IPCC report (1996) states that the climate changed over the last 100 years, and maintains that the average global air temperature on land surface

366 | Joaquín Espinosa, Sara Silva-Salvado, and Óscar García-Martín increased by 0.3–0.6 °C since the 19th century, with the most recent years being the warmest since 1860, despite the cooling effect from the volcanic eruption of Mount Pinatubo in 1991. The third IPCC report (2001) includes a larger volume of studies on climate change and, consequently, it describes a clear picture of global warming on the planet and other climatic changes as well. This report holds that the average global temperature increased since 1861 and determines that over the twentieth century this increase was 0.6 °C; however, it specifies that other important climate aspects appear not to have changed. In the fourth IPCC report (AR4, 2007), the statements regarding the effects of climate change become stronger, stating that the warming of the climate system is unequivocal,

as has been evidenced from numerous observations of the increased global average of land surface air temperature and ocean surface temperature, widespread disappearance of snow and ice, and the rise of the oceans’ global average level. This report further states that eleven of the last 12 years (1995–2006) rank among the 12 warmest years since there are global records of the land surface air temperature (1850). The update of the average increase in temperature for the last 100 years (period 1905–2005) provides a value of 0.74 °C (range 0.56–0.92 °C), being higher than for the preceding 1901–2000 period analyzed in the third Report (TAR) where it was 0.6 °C (range 0.4–0.8 °C). In the Summary for Policymakers of the Synthesis Report of 2014 [67] it can be read literally: Warming of the climate system is unequivocal, and since the 1950s many of the observed changes are unprecedented over decades to millennia. The atmosphere and ocean have warmed, the amounts of snow and ice have diminished, and sea level has risen.

This Synthesis Report also shows that each of the past three decades has been successively warmer on the terrestrial surface than any preceding decade since 1850. The period 1983–2012 was probably (66–100 % confidence) the warmest 30-year period in the last 1400 years in the Northern Hemisphere, where such an assertion is possible. The global averaged combined land and ocean surface temperatures show a warming of 0.85 °C (range 0.65–1.06 °C) over the period 1880–2012; the results were obtained using data sets from multiple independent sources [67]. In summary, when observing the evolution of successive IPCC reports about the mean air temperature on the land surface, there is absolute certainty that climate change is a reality. Regarding the oceans’ level, something similar occurs. The first IPCC report (FAR, 1990) indicated that over the last 100 years the average sea level rose in 10–20 cm, although that increase was neither linear in time nor uniform in all the oceans of the globe. The second IPCC (1996) report states that the sea level rose, on average, between 10 and 25 cm over the past 100 years, attributing most of that increase to the overall effects caused by the global temperature rise. The third IPCC report (TAR, 2001) indi-

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cates that, based on the measurements of the tidal amplitude, it is estimated that over the twentieth century global average rise in sea level is in the range of 10–20 cm. The fourth IPCC report (AR4, 2007) specifies that the global average ocean sea level increased at an average rate of 1.8 mm/year (range: 1.3–2.3 mm/year) over the period 1961–2003, with the most rapid rate in the period 1993–2003 when the value was 3.1 mm/year (2.4–3.8 mm/year). The total estimated increase over the twentieth century was 0.17 m (range 0.12–0.22 m). The Synthesis Report (2014) [67] indicates that over the period 1901–2010, the global average of ocean sea level rose by 0.19 m (range 0.17–0.21 m), and that the rate of increase in the mean sea level since the mid-nineteenth century was higher than the average rate over the past two millennia. The Synthesis Report also notes that for the period of 1992–2011, the area of Greenland and Antarctica that is covered in ice has lost mass. This loss was more intense and marked over the period from 2002–2011; glaciers continue to shrink almost worldwide as well. The spring snow cover in the Northern Hemisphere has continued to decline in extent, and there is high confidence that permafrost temperatures have increased in most regions since the early 1980s in response to increased surface temperatures and reduced snow cover. This report also states that the annual mean extent of Arctic sea ice decreased during the period 1979–2012 at a rate in the range of 3.5–4.1 % per decade. The Arctic Ocean ice-cover extension decreased each season and each successive decade since 1979, with the fastest decline in the average extension per decade occurring in the summer. It is very likely (90–100 % probability) that the annual average extension of the Antarctic sea ice cover had increased in the range of 1.2–1.8 % per decade between 1979–2012. However, there is broad consensus among researchers about the existence of strong regional differences in Antarctica, with the ice-cover extension increasing in some areas and decreasing in others. From a paleoclimatic perspective, it can be said that climate varies naturally in all time scales, either by choosing periods of hundreds of millions of years or from year to year [49]; and in the earth’s history closest to the human beings’ development, the period of glacial-interglacial cycles started about 100 000 years ago when Earth’s climate was much colder than today. During such cycles, the average global surface temperature probably varied between 5–7 °C – with major changes in ice-volume and sea level – reaching values of 10–15 °C in some mid- and high-latitudes of the Northern Hemisphere. Since the end of the last ice age about 10 000 years ago, global land surface temperatures fluctuated less than 1 °C, with some fluctuations lasting several hundred years, including the so-called Little Ice Age that ended in the nineteenth century and that seems to have been global in extent. Besides this fact, paleoclimatic data also supports the interpretation that warming over the last 50 years has been unusual in at least the past 1300 years [49]. The main factor causing disturbances in the earth’s climate is the accumulation of greenhouse gases (GHG) in the atmosphere [67]. The Synthesis Report (2014) indicates that anthropogenic emissions of GHG increased progressively since preindustrial times, mainly due to economic activities and population growth, and now

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are higher than ever. This has led to atmospheric concentrations of carbon dioxide, methane and nitrous oxide that are unprecedented in at least the past 800 000 years; their effects, along with other anthropogenic factors, have been detected through the climate system and it is extremely likely (95–100 % probability) that they are the dominant cause of global warming observed since the mid-twentieth century [67]. Between 1750 and 2011 anthropogenic CO2 emissions accumulated in the atmosphere were 2040 ± 310 gigatons of CO2 (Gt CO2 ) (1 Gt = 109 tons), and about 40 % of these emissions remained (880 ± 35 Gt CO2 ); the rest was removed from and stored in the land (plants and soil) and in the oceans. The oceans absorbed about 30 % of the CO2 emitted, causing their acidification. About half of the anthropogenic CO2 emissions between 1750 and 2011 took place over the last 40 years. Anthropogenic GHG emissions have been continuing to increase since 1970 to 2010, with the largest increases being between 2000 and 2010, despite the implementation of a growing number of mitigation policies for the climate change; and the anthropogenic GHG-emissions in 2010 reached a value of 49±4.5 Gt CO2 . CO2 emissions generated by fossil fuel consumption and industrial processes account for approximately 78 % of total GHG emissions from 1970 to 2010, with a similar percentage of contribution to the increase from 2000 to 2010 [67]. Globally, economical activities and population growth continue to be drivers of the CO2 emissions increases owing to the fossil fuel consumption. The contribution of population growth between 2000 and 2010 remains virtually identical to what happened in the past three decades; however, the contribution of economic growth to the GHG emissions has increased strongly, driven by the increased use of coal for the global energy supply. The evidence supporting human influence on the climate system has increased since the publication of the Fourth IPCC-Report (AR4) [47]. It is extremely likely (95–100 % probability) that more than half of the observed increase in the global temperature average of the earth’s surface from 1951 to 2010 was caused by the increased anthropogenic GHG concentrations, along with other forces with an anthropogenic origin. Anthropogenic forces probably contributed (66–100 % probability) in increasing the surface temperature of the planet since the mid-twentieth century in all continental regions except Antarctica. Probably, anthropogenic-forced influences have affected the earth’s global water cycle since 1960, and have also increased the rate of ice cover disappearance in Greenland since 1993. These anthropogenic influences have very likely (90–100 % probability) and substantially contributed to: (a) the loss of Arctic ice since 1979; (b) an increase in the heat content of ocean’s surface-layer (0–700 m depth); and (c) the global increase of the average sea level observed since 1970 [67].

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11.2.3.2 Projected changes in the climate system The following is written in the IPCC Synthesis Report (2014) [67]: Surface temperature is projected to rise over the 21st century under all assessed emission scenarios. It is very likely that heat waves will occur more often and last longer, and that extreme precipitation events will become more intense and frequent in many regions. The ocean will continue to warm and acidify, and global mean sea level to rise.

These changes are expected in the years 2081–2100 as compared to the period 1986– 2005. Projections of changes in the climate system are elaborated using a hierarchy of climate models ranging from simple climate models to other integrated, thorough models of intermediate complexity, as well as earth system models [48]. To construct projections of climate-change information on future emissions or concentrations of GHG, aerosols and other climate change drivers are needed. In the IPCC Fifth Report (AR5) [55], for the 21st century the scientific community has defined four emission scenarios known as Representative Concentration Pathways (RCPs) [48, 49, 67, 72–74] describing four different routes of emissions of GHG that would reach determined concentrations in the atmosphere in 2100, as well as emissions of other air pollutants (short-lived gases and aerosols) and different land uses. The RCPs include a stringent mitigation scenario (RCP 2.6), two intermediate scenarios (RCP 4.5 and RCP 6.0) and a scenario with very high GHG emissions (RCP 8.5). These scenarios, unlike those used in previous IPCC-reports (e.g. see the Special Report on Emission Scenarios, SRES-scenarios, used in TAR and AR4) [75], have been selected from various published scientific works on the subject [76–81]. Each and every one of the four scenarios is identified by the values of the radiative forcing (RF) – calculated by different authors – representing the value of the RF-stabilization in 2100. Thus the RCP 2.6 indicates that throughout the twentyfirst century the RF-value might reach a peak of 3.0 Watts/m2 (W m−2 ), but in 2100 this value will have dropped to 2.6 W m−2 . Radiative forcing (RF) is a measure of the net change in the energy balance of Earth in response to an external perturbation. Changes in the atmosphere, land, ocean, biosphere and cryosphere (both natural and anthropogenic) can perturb the Earth’s radiation budget, producing a radiative forcing (RF) that affects climate. RF is expressed in Watts/m2 (W m−2 ) [49, 82]. The RCP span encompasses the full range of RF associated with the emission scenarios published in peer-reviewed literature existing at the time of development of these RCPs, and the two intermediate scenarios were chosen because they were roughly equally spaced from the two extreme scenarios (2.6 and 8.5 W m−2 ). In sum, the land-use scenarios of RCPs show a wide range of possible futures, ranking from a net reforestation to further deforestation, consistent with projections in the full scenario literature. For air pollutants such as SO2 , the RCP scenarios assume a consistent decrease in emissions as a consequence of assumed air pollution control and GHG mitigation policy. Importantly, these future scenarios do not account for possible changes in natural forcings (e.g. volcanic eruptions) [67].

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The future of the earth’s climate system will depend on the warming caused by anthropogenic emissions in the past and those occurring in the future, as well as natural climate variability. The overall change in average of surface-temperature for the period 2016–2035, in relation to the 1986–2005 period, is in the range of 0.3 to 0.7 °C (medium confidence level), and is similar in the four RCP scenarios. This assessment is based on the assumption that no major volcanic eruptions or persistent changes in total solar irradiance will occur. Projections point that it is likely that the increase in global mean surface temperature for the period 2081–2100, in relation to the 1988–2005 period, will be placed in the intervals of 0.3 to 1.7 °C (RCP 2.6), 1.1 to 2.6 °C (RCP 4.5), 1.4 to 3.1 °C (RCP 6.0) and 2.6 to 4.8 °C (RCP 8.5). The Arctic region will warm faster than the global mean and the mean warming in continental areas will affect them to a greater extent than over the oceans (very high confidence level). It is likely that by the end of the twentyfirst century, global surface temperature will increase by 1.5 °C when compared to the period between 1850 and 1900 for all RCPscenarios, except for the RCP 2.6. It is also likely that such temperature will exceed 2 °C for both RCP 6.0 and RCP 8.5 scenarios, and more likely than not that it will exceed 2 °C for the RCP 4.5 scenario. The warming will continue after 2100 in all RCP scenarios, except for the RCP 2.6; it will continue to display variability between interannual and decadal, and it will not be uniform across regions [48, 67]. The global ocean will continue to warm throughout the twentyfirst century, with a stronger warming of the sea surface expected for the tropics and subtropical regions of the Northern Hemisphere. Projections point to a global increase in ocean acidification for all RCP scenarios at the end of the century. The decrease in pH of ocean surface will be in the range of 0.06 to 0.07 pH units (15–17 % acidity increase) for RCP 2.6 scenario, 0.14 to 0.15 pH units (38–41 %) for the RCP 4.5 scenario, from 0.2 to 0.21 (58–62 %) for the RCP 6.0 scenario, and from 0.30 to 0.32 (100–109 %) for the RCP 8.5 scenario. Since the 4th IPCC-report (AR4) was published [71], progress has been significant in understanding and projecting changes in sea level. The mean global rise of sea level will continue throughout the century, most likely at a higher rate than observed between 1971 and 2010. For the period 2081–2100, in relation to the period 1986–2005, such an increase will likely be found in the range of 0.26 to 0.55 m for the RCP 2.6 scenario and from 0.45 to 0.82 m for the RCP 8.5 scenario (medium confidence level), and this increase will not be uniform in all regions. At the end of twentyfirst century it is very likely that sea level will have risen by more than 95 % of all the oceanic area, and about 70 % of coastlines around the world will have undergone a sea level rise of at least 20 % of the global mean [67]. Cumulative CO2 emissions will largely determine the global average of surface warming at the end of the twentyfirst century and beyond. Most aspects of climate change will endure for many centuries, even though CO2 emissions would stop, which is a remarkable inevitability of climate change for centuries, because of CO2 -emissions

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in the past, present and future; the risks of abrupt and irreversible changes will increase as the magnitude of warming increases. A large part of anthropogenic climate change resulting from CO2 emissions is irreversible on a time scale of several centuries to millennia, except that an abundant net removal of CO2 from the atmosphere for an extended period occurs. For many centuries, the surface temperature will remain approximately constant at high levels after complete cessation of net anthropogenic CO2 emissions. Given the dilated temporal scales at which heat transfer occurs from the surface to the deep ocean, ocean warming will continue for centuries. Depending on the emissions scenario considered, approximately between 15 and 40 % of the CO2 emitted will remain in the atmosphere for a period longer than 1000 years [48].

11.2.3.3 Major effects of global climate change on the oceans General aspects The ocean is a vast region extending from the high tide line to the deepest ocean trench (11 030 m), occupying 71 % of the earth’ surface. The total volume of the ocean is about 1.3 billion km3 , with about 72 % of this volume being below 1000 m (what is known as the deep sea) [65]. The oceans produce half the oxygen we use to breathe and to burn fossil fuels, and provide about 17 % of the animal protein consumed by the entire world population, or almost 20 % of the protein consumed by 3000 million people. In the oceans, countless species and their ecosystems have great value in tourist and recreational activities. Ocean ecosystems such as coral reefs and mangroves protect coastlines from tsunamis and storms. About 90 % of the goods that the world uses and shares travel through sea routes with ships ploughing the oceans. All these activities will be affected by the climate change [64]. The ocean plays a major role in the dynamics of global climate, absorbing 93 % of the accumulated heat in the atmosphere, and the consequent heating of the oceans affects most ecosystems. Around a quarter of all CO2 emitted from the consumption of fossil fuels is absorbed by the oceans. The oceanic plankton transforms a part of this CO2 into organic matter, part of which is exported to deeper ocean zones. The remaining CO2 causes a progressive ocean acidification from chemical reactions between CO2 and seawater, and this acidification is enhanced by both the supply of nutrients and the intensification of the loss in oxygen content. All these changes involve risks to life in the oceans and can affect its ability to perform the wide range of functions that are of vital importance to human health and the environment [64]. The effects of climate change take place in an environment that also has natural variability in many of these variables. Other human activities also influence the conditions of the oceans, such as overfishing, pollution and nutrient supply through the rivers; this causes eutrophication, a process that generates large areas of water with very low oxygen content (called dead areas). Theis wide range of factors affecting ocean conditions and the complexity with which they interact make it difficult to

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identify the role played by a single factor in the context of climate change, while they also complicate the studies to identify precisely the combined effects of these multiple drivers of climate change [64].

Changes in physical, chemical and biological variables of the ocean Trends in ocean conditions over the past 60 years reflect significant human impacts, which going beyond natural variability, on temperature, salinity, dissolved inorganic carbon, oxygen content, pH and other properties of the upper ocean [64, 83, 84]. With climate change in the future, marine ecosystems are and will be exposed to increases in temperature, acidification processes, expansion of hypoxic zones and other drivers of environmental changes acting simultaneously.

Temperature and salinity Over the period from 1971 to 2010, the oceans have warmed at an average rate of more than 0.1 °C per decade in the surface zone (first 75 m) and 0.015 °C per decade at a depth of 700 m. Even during this period, it is likely that the intensification of this warming near the surface has caused thermal stratification of the upper layers of the ocean around 4 % between 0 and 200 m in depth in all parts of the ocean at north of 40° south latitude (40° S). It is also likely that the ocean has warmed between 700 and 2000 m from 1957 to 2010; warming signals are less apparent or nonexistent at greater depths. These changes include a virtually certain anthropogenic signal [65, 83, 85], with the water surface from the three major ocean basins (Indian, Atlantic and Pacific) warming at different rates but exceeding the ones expected if there were no changes in GHG concentrations driving those alterations over the past century [65]. It is very likely that the superficial layers of the three major ocean basins have warmed, with the Indian Ocean (0.11 °C warming average per decade) warming faster than the Atlantic (0.07 °C per decade) and the Pacific (0.05 °C per decade). These data are consistent with the average of temperature variation in depths from 0 to 700 m observed in the period from 1971 to 2010. During the period 1950–2009, while some oceanic zones (e.g. the North Pacific) did not show a clear warming trend, most oceanic regions showed either significant warming in mean temperature, or a significant warming in either/or the warmest and coolest months of the year. Trends in the temperature of the sea surface show a considerable sub-regional variability. Notably, the average temperature of most regions called “Northern Hemisphere High-Latitude Spring Bloom Systems (HLSBS)” did not show significant increases in temperature over the period from 1950 to 2009 (except in the Indian Ocean); although the temperatures of the warmest month of the year in the North-Atlantic and South-Atlantic, as well as in the Southeast-Pacific and the temperatures of the coldest month of the year in North and South Atlantic and South Pacific, showed a significant upward trend during this same time period (1950–2009) [65].

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The global water cycle on earth is dominated by the processes of evaporation and precipitation occurring in oceans, and the sea surface salinity varies with temperature, solar radiation, cloud cover and ocean circulation [65, 86]. The ocean plays a central role in the global water cycle on earth: about 85 % of evaporation and 77 % of the precipitation occurs over the ocean [51, 87]. The horizontal distribution of salinity in the upper ocean zone mostly reflects this exchange of fresh water, with high salinity values in regions where evaporation exceeds precipitation, and low salinity in zones where rainfall and the input of fresh water from rivers or runoffs predominate; oceanic circulation also affects the regional distribution of surface salinity. The subduction of surface water transfers the surface salinity into the ocean, so that the distribution of salinity in the areas below the ocean surface (subsurface salinity) will also depend on the patterns of evaporation, rainfall and inland freshwater input to the sea surface. The melting and formation of both sea ice and glacial ice also influence ocean salinity. Trends in salinity values are consistent with the amplification of the global hydrological cycle [65, 83, 88] as a result of a warmer atmosphere, which is most likely the origin of trends observed towards greater precipitation, evaporation, humidity of the atmosphere and extreme meteorological events. The spatial patterns in salinity and evaporation/precipitation are correlated, which is clear indirect evidence that these processes have experienced an increase since the 1950s. It can be expected that the water cycle on the earth intensifies in a scenario of a warmer climate because the warm air can make the climate more wet: the atmosphere may increase its content in water vapor by 7 % per 1 °C increase in its temperature. Observations made since 1970 [51, 89] show increases in the amount of water vapor in the lower layers of the atmosphere, which is considered consistent with the observed warming. It is also predicted that in a warmer climate evaporation and precipitation will intensify; changes in salinity records in the oceans over the past 50 years support these predictions. Seawater contains both salt and fresh water, so that salinity is a function of the weight of salts dissolved in seawater. Because the total amount of salt coming from the weathering of rocks does not change during human timescales, the salinity of seawater can only be altered (in days or in centuries) by adding or removing fresh water. The atmosphere connects regions of the ocean where a net loss of water occurs with those where the opposite happens by moving the evaporated water from place to place; therefore, the distribution of salinity in the ocean surface reflects mostly the following: a spatial pattern of evaporation minus precipitation and terrestrial water inputs and processes related to sea ice. The subtropical oceanic waters show higher salt concentration because evaporation exceeds precipitation, while both tropical oceanic regions and high oceanic latitudes – where rainfall exceeds evaporation – are less saline. The Atlantic, the saltier oceanic basin, loses more water through evaporation than it gains in precipitation; while in the Pacific the balance is almost neutral and in the Southern Ocean (region around Antarctica) precipitation predominates.

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The salinity changes in recent years in the sea surface and upper ocean (0–700 m depth) reinforce the average salt-concentration pattern: the oceanic subtropical regions, where evaporation dominates, have become saltier; while both tropical and subpolar regions, dominated by precipitation, have become sweeter. When salinity changes are measured in the more superficial 500 m of the ocean, these trends persist but become even more pronounced. The global and direct observation of changes in precipitation and evaporation is difficult because most of the freshwater exchange between the atmosphere and earth surface occurs over the 70 % of the earth surface covered by oceans; furthermore, rainfall records over long periods of time are only available for the earth’s surface, where, additionally, there are no prolonged evaporation records. Rainfall observations made at the terrestrial surface in recent years show increased precipitation in some regions and a decrease in others, making it difficult to be able to build a global and integrated picture of the phenomenon. These observations have revealed episodes of extreme rainfall and flooding associated with the fact that the winter snow, in northern latitudes, melted prior and in coincidence with times of the year where heavy rains still occur; however, these observations still have a strong regional component in their trends. Therefore, precipitation observations made at the terrestrial surface are, so far, insufficient to provide evidence of changes in drought areas. Moreover, ocean salinity acts as an effective and sensitive rain gauge, reflecting and evidencing the differences between water gained by the ocean, by precipitation and water lost through evaporation – both random and episodic events. The oceanic salinity is also affected by the water supply from the continents and freeze/thawing of both glacial ice and the sea ice floating in the ocean. The input of fresh water from melting ice or terrestrial runoff modifies the global salinity average, but, to date, the changes are not important enough for them to be observed [51].

Concentrations of CO2 and acidification The increase in the concentration of carbon dioxide (CO2 ) in the air, measured as partial pressure of CO2 (pCO2 ) and expressed as μatm – also equivalent to the expression “parts per million” or ppm – causes an increase in the CO2 concentration in the upper strip of the ocean [64, 90]. Starting from a pre-industrial value of 280 μatm in the atmospheric CO2 concentration, by 2050 values of about 500 μatm [75] are expected for all the defined emission scenarios (RCPs) [73, 82]. By 2100 it is anticipated that the atmospheric CO2 concentration could reach values in the range of 420–940 μatm depending on the emission scenario (RCP). The increase in the partial pressure of CO2 in the atmosphere causes ocean acidification – measured as the decrease in pH of the water – and this acidification is accompanied by a decrease in both the concentration of carbonate ion (CO2− 3 ) and the saturation status of various calcium carbonates (e.g. CaCO3 ) [91], while the bicarbonate ion (HCO−3 ) increases [51].

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The variations in the total dissolved inorganic carbon in ocean water (CT = CO2 + − CO2− 3 + HCO3 ), as well as the variations in the partial pressure of CO2 (pCO2 ), reflect changes in both the natural carbon cycle and the amount of CO2 that the ocean fixes from anthropogenic CO2 emissions to the atmosphere. In the open ocean, the pH average of the surface water is in a range from 7.8 to 8.4; therefore the ocean pH-average is currently slightly alkaline [92]. The decrease by 0.1 pH (from 8.2 to 8.1) in the pH average of ocean surface waters from the beginning of the industrial age correlates with an increase of 26 % in the concentration of hydrogen ions [H+ ] in sea water [92]. Estimates for future concentrations of atmospheric CO2 and in ocean water predict that by the end of the twentyfirst century (2100), the pH average of ocean surface water will be lower than it was over the past 50 million years [93]. The main parameters controlling pH variations of ocean water are: (a) the exchange of CO2 with the atmosphere; (b) the production and respiration of particulate organic matter and dissolved in the water column; and (c) the formation and dissolution of calcium carbonate minerals. The oxidation of organic matter in seawater decreases the oxygen concentration, adds CO2 to water and lowers the water pH and saturation concentrations of both carbonate ions (CO2− 3 ) and calcium carbonate (CaCO3 ) [94]. As a result of these processes, the minimum pH values of ocean waters are generally found close to, or coinciding with, zones of minimum oxygen concentrations (Oxygen Minimum Zones, OMZ). When CO2 reacts with seawater, it forms carbonic acid (H2 CO3 ), a highly reactive molecule that reduces the concentration of carbonate ions (CO2− 3 ), which might affect the process of shell formation of a large amount of marine organisms, such as, among others, corals, some species of plankton and shellfish. This process might affect biological and chemical mechanisms of large number of marine organisms in the coming decades [95]. The species of marine organisms that use calcium carbonate (CaCO3 ) to build their skeletons or shells are called calcifier organisms. Sea water contains large amounts of calcium, but for the calcifier organisms to use it in the form of calcium carbonate, such organisms need to enable specific body compartments to increase the alkalinity of solutions far above other body compartments, and far above seawater itself as well, and this process requires energy. If the sea water entering an organism has high CO2 levels, it may alter the internal body acidity values; thus maintaining alkalinity values needed for processes such as shell production by a mollusk will be far more costly for the organism from an energetic point of view. Obviously, the more energy must be devoted to the calcification processes, the less is available for other biological processes such as growth and reproduction, and therefore it can lead to a reduction in the viability of the organism as such. The response of different marine calcifier organisms to the ocean’s progressive acidification is diverse; they don’t all have the same sensitivity, but bivalve mollusks are among those with a more pronounced response to this parameter . In fact, the accused sensitivity with which bivalve mollusks respond

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to this variable of acidification remind one of the already proven response that these organisms experienced during the Permian mass extinction [96, 97].

Oxygen concentrations: hypoxia and anoxia The amount of dissolved oxygen in seawater is one of the main conditions for marine organisms’ distribution and abundance. Oxygen concentrations in seawater vary across different ocean basins, being lower in the Eastern Pacific, Atlantic and northern Indian Ocean. In contrast, some of the highest oxygen concentrations are associated with the cold waters of higher latitudes [65]. The average concentration of dissolved oxygen in the ocean is currently (2006) 162 μmol kg−1 [98]. These oxygen concentrations span a range from more than 500 μmol kg−1 in oxygen-supersaturated Antarctic waters with high productivity [99] to zero-concentration zones (anoxic zones) in coastal areas with high concentrations of sediment, as well as in deep layers of permanently anoxic isolated waterbodies, such as the Black Sea or the Gulf of Cariaco. Hypoxia is the result of an excessive consumption of oxygen that greatly exceeds the normal supply of it, as it occurs in highly stratified marine areas. Wide zones with Oxygen Minimum Content (Oxygen Minimum Zones, OMZs) are found between depths of less than 100 m and those over 900 m in the eastern Atlantic and in the tropical Pacific Ocean. In ecological literature, hypoxia – or hypoxic conditions – is defined as waters containing oxygen concentrations below 60 μmol kg−1 , estimating that these zones correspond to around 5 % of global ocean volume [100]. The Pacific OMZ zones typically have oxygen concentrations below 20 μmol kg−1 , representing about 0.8 % of overall volume of the ocean [101]; these concentrations are lower than those of the OMZ zones of the Atlantic Ocean. Waters with very low oxygen concentrations (suboxic waters), with values of less than 4.5 μmol kg−1 occupy about 0.03 % of the ocean volume, mainly in the tropical Pacific Northwest [102]. The OMZ-zones are naturally located in marine habitats with high sediment concentration, but they are also expanding because of anthropogenic factors. Over the past 50 years, the oxygen concentrations in the open ocean have fallen to an average between 0.1 to more than 0.3 μmol kg−1 per year [103]; in some OMZs the rate of decrease in oxygen concentration was much higher due to warming, the increasing stratification of ocean layers and the increasing of oxygen biological demand. It has been reported that oxygen concentration decreases about 7 μmol kg−1 per decade at average depths in many zones of the subarctic North Pacific [104]. In 48 years (1960– 2008), bibliographic references describing extremely hypoxic “dead zones” in coastal regions have increased almost tenfold; the causes of this increase in OMZ-zones are all of anthropogenic origin: high oxygen demand because of eutrophication, resulting in a high load of organic matter whose decomposition consumes oxygen from water, and also the formation and release of nitrous oxide [105–107]. Because of the warming the

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earth is experiencing, it is expected that hypoxic zones will spread across the oceans, especially in both temperate and sub-polar areas [64]. Most climate models based on the RCP 8.5 scenario predict a decrease in the global ocean oxygen concentration from 1 to 7 % between now and 2100 [104], and with an average decrease of 3.4 % in 2090 when compared to the ocean oxygen concentration in 1990 [108]. It is considered that the evolution of regions with low O2 concentration may be influenced by changes in the inputs of inland water [109] and wind regime [110], as well as by the intensity, duration and seasonality of the different upwelling systems [111]. Climate change may influence the distribution of ocean zones with low or very low oxygen concentrations (dead zones) by increasing the ocean water temperature and microbial activity, as well as by reducing the mixing effect between different zones of the ocean (e.g. by increasing the stratification process), and therefore reducing the incursion of oxygen-rich ocean surface layers into deeper ocean zones. In other ocean zones, the increase of upwelling phenomena could lead to a stimulation of the trophic chain productivity, which in turn would cause a greater amount of organic matter to reach the deep ocean where it is consumed, but causing a decrease in oxygen concentrations. By controlling local factors such as the nutrients supply to coastal regions, it could be achieved to reduce the speed to which the dead zones expand across the oceans [65]. Anyway, currently there is still no broad consensus on the future evolution of the volume to be reached by hypoxic and suboxic waters, mainly due to strong uncertainties about potential biogeochemical effects and the evolution of ocean dynamics under the influence of possible alterations, both natural and anthropogenic [64].

Light, nutrients and primary production Most of the existing climate models currently predict that in a near future the mixingzone of the ocean surface – which includes the totality of the euphotic zone, i.e. that zone where the light intensity allows the realization of some photosynthetic activity and that usually not exceed 300 m in depth [112] – will become more shallower mainly due to the strengthening of the vertical density gradient [64, 113, 114]. Mean levels of light that phytoplankton communities find are determined by these variables: light incident on the sea surface from solar radiation, depth of the mixing- one and the degree of light attenuation below the seawater produced by particles existing there [115]. A reinforced mixing zone will result in phytoplankton receiving a greater amount of light whenever the distribution of organisms is more or less uniform. The increasing ocean stratification [116] – particularly in certain regions such as the North Atlantic, Northeast Pacific and Arctic [117] – will cause a decrease in the vertical transport of nutrients from the deep ocean to more superficial waters [118], which will impact on photosynthetic organisms living in these superficial waters. This nutrient decrease may be offset, at least in part, because of the eutrophication of human

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origin and increased upwelling phenomena in certain areas, and by the mix caused by tidal phenomena and the estuarine water circulation in coastal zones [64]. Ocean net primary production (NPP) is the amount of carbon fixed by photosynthetic organisms living in the upper ocean zones [60, 64]. The availability of light and nutrients for photoautotroph organisms determines, daily, the net primary productivity; this availability may be altered, directly or indirectly, by modifying the ocean mixing zone depth, changing ocean circulation regimes in different ways, or provoking the physical displacement of organisms. A climate change will affect the mixing zone depth, the cloudiness over the ocean (increase) and the amount and thickness of floating ice in the ocean; all of these effects will impact the NPP [64]. A reinforcement of the vertical density gradient in the ocean surface zones will reduce the communication between the ocean euphotic layer – bathed by sunlight and where the photosynthetic activity takes place – and the ocean’s lower zone rich in nutrients. The supply of macro-nutrients (such as nitrate) and micro-nutrients (such as iron) to marine plants [119] varies both seasonally [120] and regionally [121], so that NPP might be simultaneously co-limited by more than one variable [122]. From a global perspective, it is expected that until the year 2100, the NPP will moderately descend in any of the three emission scenarios: low (RCP 4.5), moderate (RCP 6.0) and high (RCP 8.5). At the same time, it is expected that the NPP will increase at high latitudes and decrease in the tropics (medium confidence level). Anyway, there is currently limited evidence and low agreement about the direction, magnitude and differences of the NPP changes projected by the year 2100 in different oceanic and coastal zones [60].

11.3 Reproductive cycle, sex expression and sex change in shellfish 11.3.1 Reproductive cycle, reproductive period and sex expression in bivalve mollusks The gonad of bivalve mollusks is a system that deploys a recurring temporal evolution where the system returns incessantly to a state proximal or like others it had in the past; consequently, the gonad performs cycles with an associated time period. We define the reproductive cycle of a bivalve mollusk as all cellular processes experienced by the gonad of the animal leading to the production and spawning of gametes, and such processes occur in a sequential and cyclic manner. The stages of this cycle are: restoration, gametogenesis (spermatogenesis, oogenesis), spawning and resting. Gamete maturation is the final phase of gametogenesis; and these stages occur sequentially in this order. We define the reproductive period of a bivalve mollusk as the temporal extent within which a reproductive cycle occurs, i.e. the duration of a reproductive cycle.

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The reproductive period lasts one year and in the northern hemisphere usually extends from the start of autumn until the end of summer in the following year. The reproductive cycle starts with the restoration stage and then continues with gametogenesis, maturation and spawning of gametes; it ends with the resting stage of the gonad. In each breeding season, the mollusk builds one gonad where, depending on the type of sexuality, the animal can develop sexual phases: male, female or both sexual phases. When a reproductive period ends, the gonad returns to the initial state and begins a new cycle to install the next reproductive cycle. Dioecious (unisexual) mollusks develop, mature and spawn in only one sexual phase – male or female – in a reproductive cycle, but ambisexual (bisexual) mollusks, depending on their type of sexuality, can grow and spawn one, two or more sexual phases of different sexes in a reproductive cycle. The reproductive cycle is seasonal, starting in autumn with the restoration stage (September–October), followed by gametogenesis and gamete maturation (rest of autumn, winter and spring), spawning (late spring and first half of summer) and ends with the resting of the gonad (second half of the summer and early autumn). This seasonality of the reproductive cycle suggests that environmental factors such as temperature and photoperiod could influence the onset and development stages of the cycle, particularly gametogenesis and spawnings. Likewise, in ambisexual species, the temperature and photoperiod could influence the timing and sequence the animals perform the sex change. Consequently climate change, with the changes in temperature and photoperiod involved, could significantly influence the development of the reproductive cycle and sex change of bivalve mollusks.

11.3.2 What is sex? In multicellular organisms that reproduce sexually, sex is a fundamental biological trait that has profound importance in the development and reproduction of individuals and in the formation of the sex ratios of natural populations [123]; consequently, sex has great significance in biological evolution [124]. The sex of an organism is a sexual phenotype arising from the genotype; its expression occurs under regulatory conditions. When an organism expresses sex, it acquires a sexual phenotype that in dioecious organisms can be male or female. Bisexual – hermaphroditic – organisms have the ability to express both sexes (male and female) in their lifetime. The origin of the sex of an organism is in its genotype (genetic sex) whose expression results in the sexual phenotype of the organism. When an organism is born, the sexual phenotype is expressed on two levels: (a) in the soma of the gonad (somatic sex); and (b) in the germline that develops inside the gonad (germline sex). The sex of an organism is determined by the gonad somatic sex and is not determined by the germline sex or by the genetic sex [125, 126]. The sex expressed by the soma of the gonad is a somatic sexual phenotype – i.e. is somatic sex – and gives the sexual iden-

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tity to the organism [125, 126]. This means that if the gonad expresses a male sex, it becomes a testicle and the organism will have a male sex; and if the gonad expresses a female sex, it becomes an ovary and the organism will have a female sex. The germline cells develop inside the gonad and express a sexual phenotype (germline sex) male or female induced by the gonad somatic sex, i.e, it is subject to the gonad somatic sex [125–140]. When genetic and somatic sexes match on a dioecious organism, it is said that the organism has a coherent sexual system. However, if genetic sex and somatic sex are different, it is said that the organism has an incoherent sexual system, and also, that it has sex reversal.

11.3.3 Sex determination: everything happens in the embryo Bringing together established knowledge on embryonic development of animal models such as Drosophila [129, 138, 140–142], Caenorhabiditis elegans [140, 142], fish [136, 143], reptiles [132, 136, 137], birds [136] and mice [126, 132, 133, 137], in the embryonic development of an organism two phenomena occur that happen and initiate independently and then converge together. These phenomena are: (a) the gonad formation and the determination of its sex (somatic sex); and (b) the formation of primordial germ cells (PGC) – which will produce the gametes – and the determination of their sex (germline sex). In the embryo, the gonad is formed from the genital ridge that phenotypically is sexually bipotential, meaning that although it has only one genotype, it can give rise to two distinct somatic sexual phenotypes in two different sexualized gonads, and each of these two phenotypes is called the somatic sex of the gonad. Each somatic sex, male or female, causes a different type of gonad. Therefore, the gonadal somatic sex is a sexualized gonad, and PGC are not involved in the processes where gonad acquires its somatic sex [126, 133, 137, 138]. From the genital ridge, two distinct gonadal sexual phenotypes can differentiate: male sexual phenotype (MSP), which causes a testicle, and female sexual phenotype (FSP), which produces an ovary; both are somatic sexual phenotypes. The genital ridge, being sexually bipotential, has to decide whether it differentiates into an MSP or FSP. It is a decision or switch between two possibilities known as primary sex determination [137]. When the genital ridge has carried out the process of sex determination, a process of differentiation to produce a MSP or a FSP begins; this process of differentiation is called the sex differentiation of the gonad [137]. This gonad sex differentiation can take one of two paths: a pathway that produces a MSP (a testicle), or a pathway that results in a FSP (an ovary). Primordial germ cells (PGCs) originate in the embryo and set up the organism germline that will produce the gametes. PGCs are sexually bipotential [130, 132, 133, 138], meaning that the same genotype can express one of two different sexual phenotypes: a male phenotype or a female phenotype. In the embryo, PGCs originate separate and simultaneously with the genital ridge formation [137], and, once it is sexually

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determined, that is, when it starts to produce a male or female gonad, PGCs colonize the just sexualized gonad that is forming. The PGCs differentiate sexually, i.e. acquire a male or female sexual phenotype when colonizing the sexualized gonad that is forming, and their sexual phenotype is induced by the gonad somatic sex [125, 126]. This means that the gonad consistently induces the sexual differentiation pathway PGCs will take to produce its sexual phenotype: if the gonad’s somatic sex is male, PGCs will develop a male sexual phenotype, and if the gonad’s somatic sex is female, PGCs will develop a female sexual phenotype. When PGCs differentiate sexually and acquire a sexual phenotype, they are known as germinal stem cells (GSC) [125, 126]. According to the somatic sex of the gonad colonized by PGCs, these cells can take two different paths to sexual differentiation: (a) the spermatogenesis pathway where the founder cell is a GSC named spermatogonia; and (b) the oogenesis pathway where the founder cell is a GSC known as oogonia. PGC differentiation to cause a GSC by either of these two pathways is known as PGC sexual differentiation [140]. In the embryo, the formation of the PGCs needs two processes: (a) a determination process known as germ cell determination [144–150], and (b) a process of sexual differentiation that is called germ cell sex-differentiation [133, 137]. The determination process occurs early in embryonic life and in this process one or more cells in the embryo become PGCs, i.e. it is the process of forming the PGCs that will constitute the germ line of the organism [148–150]; in this process the gonad is not involved because it is not yet formed in the embryo, and even the genital ridge is not formed at all [148, 150]. This process of determining the PGCs can occur in two ways: predetermination or epigenesis [148]. When PGCs are formed to produce the organism germ line, these cells acquire a sexual phenotype – male or female – through a process called sexual differentiation and, as mentioned above, the sexual phenotype of PGCs is coherently induced by the somatic sex of the gonad these cells colonize [125, 126]. In short, the sexual identity of a multicellular organism is given by the somatic sex of its gonad, and when the organism is born it is endowed with: (a) a genotype; (b) a gonad somatic sex giving the sexual identity to the organism; and (c) a germ line – consisting of GSCs with a defined sexual phenotype, male or female – that will produce the gametes. In the male, this germ line consists of spermatogonia, and in the female, it consists of oogonia (non-mammalian organisms) or primary oocytes (mammals).

11.3.4 Sex determination of the gonad and sex differentiation of primordial germ cells (PGCs): molecular basis and regulation Bearing genotypic basis, the sex of an organism is fixed by the somatic sex of its gonad, which, in turn, is determined in the embryo by chromosomal factors (genotypic sex determination [GSD]) or by environmental factors (environmental sex determination [ESD]) [151–164]. In GSD species, sex is determined just at oocyte fertilization through

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a combination of sex chromosomes (mammals, birds and most fish) or sex chromosomes and autosomes (insects) [156, 159]; but in amphibians, reptiles and certain fish species, sex is determined after oocyte fertilization – during early embryonic development – by environmental factors, particularly temperature (temperature-dependent sex determination [TSD]) [151–155]. At present, three mechanisms determining the gonad somatic sex during embryo development are certainly known: (a) the genetic system SRY on the Y chromosome (mammals); (b) the genetic system DMRT1 (birds and certain fish); and (c) the estrogen-dependent system (17β-estradiol [E2]) [157]. The SRY and DMRT1 systems are found in GSD species and the estrogen-dependent system is in TSD species [157]. However, some species of fish (case of tilapia), as GSD species with genetic sex based on the chromosome pairs XY/XX, do not use the system SRY to determine the gonad’s somatic sex; instead, it depends on the estrogen (E2) levels in the embryo [165]. The presence of the SRY or DMRT1 systems in the genotype of an embryo determines that it develops a testicle, and their absence determines that it develops an ovary; this occurs in a manner independent of temperature [157, 160]. TSD species and some GSD species have no SRY or DMRT1 systems, and the determination of the gonad somatic sex in these species depends on the presence/absence of 17β-estradiol (E2) levels in the embryo: high concentration of E2 in the embryo at the time of gonad development produces an ovary, and a low concentration results in a testicle [153–155]; and in these species the regulation of E2 concentration depends on the environmental temperature [124, 151–155, 158–162]. Consequently, it is expected that changes in ambient temperature affect the determination of the gonad somatic sex more in species with an E2-dependent system than in species that have a SRY- or DMRT1-dependent system [160–164]. In the embryo, processes of sex determination and sexual differentiation of the gonad, as well as processes of determination and sexual differentiation of primordial germ cells (PGCs), depend on cascades of genes whose expression are activated/deactivated by a network of transcription factors acting on regulatory genes [157, 158, 160–163]. These transcription factors are proteins that regulate gene activity through processes of protein-gene interactions, and these processes can be affected by temperature. Consequently, the temperature may affect the sex determination and sexual differentiation of the gonad, as well as the determination and sexual differentiation of PGCs.

11.3.5 Gonad somatic sex and germline sex in bivalve mollusks Bivalve species of great commercial importance accumulate marine toxins with a high level of toxicity to humans, particularly species belonging to the families Pectinidae, Mytilidae, Veneridae, Cardiidae, Solenidae and Pharidae [166]. In these species, it is important to know how changes in environmental temperature can affect: (a) both go-

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nad somatic sex determination and gonad sexual differentiation in the embryo; (b) the determination of primordial germ cells in the embryo; (c) the sexual differentiation of primordial germ cells in the embryo; (d) the processes of spermatogenesis and oogenesis in the adult organism; (e) the reproductive cycle of the adult organism; and (f) the sex ratios in natural and farmed populations. All these aspects have a cellular basis which, in turn, relies on the action of proteins on the expression/repression of genes as well as on other cellular processes. If changes in the environmental temperature affect these processes, then the target molecule for the temperature might be the proteins involved in them. The concept of gonad somatic sex is a recent concept that has emerged from studies on Drosophila and Caenorhabditis elegans [129, 138, 140–142] and has been extended and has fertilized investigations conducted on animals vertebrates [126, 132, 133, 136, 137, 140, 142, 143]. However, this concept is not yet used when the gonad of bivalve mollusks is studied. Studies addressing different aspects of this organ – histology, physiology, biochemistry and gonadal cycle – in bivalve mollusks deal with the germline component of the gonad (germ cells, spermatogenesis and oogenesis) but not on the gonad somatic component, and even this somatic component remains to be defined. Consequently, the concept of somatic sex is not used when studying the gonad of bivalve mollusks, despite the great importance it has both in determining the sex of the organism and in the sexual differentiation of primordial germ cells. The gonad of a bivalve mollusk consists of a tridimensional network of gonoducts ending in follicles, with the gonadal follicle being the functional unit of the gonad; inside these follicles the processes of spermatogenesis and oogenesis develop that result in the production of gametes. The cellular structure of a follicle consists of a thin layer of fibroblasts constituting the follicle wall and germinal cells, as well as a few others cell types – whose functions are yet unknown – enclosed within it. From a structural and functional point of view, we define the gonad of a bivalve mollusk as consisting of two components: a somatic component and a germline component. The cellularity of the somatic component consists of a thin layer of fibroblasts forming the wall of the follicles and, also, non-germinal cells observed inside the follicles. The cellularity of the germline component consists of all the cells constituting the germline, including the gametes. Except the extensive and deep research published by Farris H. Woods in 1931 on the bivalve mollusk Sphaerium striatinum [167], little more is known about early embryonic development in bivalve mollusks [168–170]. In this study, Woods reported on the formation of gonoducts and primordial germ cells (PGC) during early embryonic development of this mollusk. Already in the embryo, when the gonad is forming, its two components appear: the somatic component and the germline component. The somatic component consists of fibroblasts forming conduits containing the germline component that, in turn, consists of primordial germ cells (PGC). Such conduits will later become gonoducts that finally form follicles in the adult animal.

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At present, no cellular and molecular processes determining both somatic sex and germline sex in bivalve mollusks has been reported. In these animals, sex is defined considering just the germline sex manifested in the gonad, i.e. the gametes’ phenotype (spermatozoa or oocytes) and the processes producing gametes (spermatogenesis or oogenesis). According to the above-mentioned knowledge from studies conducted in Drosophila, Caenorhabiditis elegans and vertebrates, we think that bivalve mollusks might have somatic sex somehow expressed in the somatic component of the gonad consisting of fibroblasts forming the follicles wall.

11.3.6 Sex, sex reversal, types of sexuality and sex change in bivalve mollusks An organism has sex reversal when its sexual system is inconsistent, i.e. when its genetic and somatic sexes are mismatched. This phenomenon can be produced experimentally during embryonic development in vertebrate TSD and GSD species wherein the determination of the gonad somatic sex is estrogen (E2) dependent [165, 171]. No such experiments have been reported to have been performed on bivalve mollusks. In vertebrates, the sex reversal phenomenon is clearly described, defined and explained [165, 171], but not so with the phenomenon of sex change; in vertebrates, except in some fish species [172], phenomena of sex change in adult individuals have not been reported. However, sex change is frequent in species of bivalve mollusks [173]. At present, a sex change is diagnosed in an adult bivalve mollusk when the animal – in its lifetime – changes from producing a gamete phenotype, e.g. spermatozoa, to produce the other, e.g. oocytes, and vice versa. Possibly, the concepts of sex reversal and sex change do not match; with current knowledge, it is known that sex reversal occurs during embryonic development of the organism but not in the adult animal; however, the sex change happens in bivalve mollusks in the adult animal but it is not known if it occurs during embryonic development as well. What needs to happen in the organism of an adult bivalve mollusk to perform sex change? This issue requires theoretical approaches and experimental studies. Vertebrates, except certain fish species, are dioecious organisms producing only a phenotype of gamete (spermatozoa or oocytes) in their lifetime, and by studying the embryonic development of vertebrate model species (mouse, and certain reptiles and fish), and also Drosophila and Caenorhabditis elegans, the concepts of sex determination, sexual differentiation of the gonad, somatic sex, determination and sexual differentiation of primordial germ cells [125, 126, 130–133, 137, 140–142, 144–148] and sex reversal [164, 165] have been developed; but bivalve mollusks species have not been employed to define or apply these concepts. Consequently, such concepts or other similar ones are not currently used in the case of bivalve mollusks. However, there is vast literature describing bivalve species which simultaneously produce both sexes or change sex in their lifetime [173], and such studies describe the processes but do not explain them at

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the cellular or molecular level. In this regard, there is still a wide range of theoretical and experimental research to be done. We define a sexual phase as the cellular expression of a germline sex – male or female – in the gonad of an organism, and this expression consists of gametes – spermatozoa or oocytes – and the cellular processes that form them, spermatogenesis or oogenesis, respectively. The gonad, in the lifetime of an organism, can express a male sexual phase (MSP) (spermatozoa and spermatogenesis), a female sexual phase (FSP) (oocytes and oogenesis) or both sexual phases. We define, at least for now, sex change as a change of the functional sexual phase in the gonad of an adult organism from a MSP to a FSP, or the reverse; therefore, an organism changes sex when it changes the functional sexual phase of its gonad; and in these terms, a sex change is defined, at least for now, as a change in the germline sex and not as a change of somatic sex. Wesley R. Coe, based upon the presence of sexual phases in the gonad, published two extensive studies, currently in force, describing (but not explaining) the types of sexuality and sex change (germline sex change, in our terms) in bivalve mollusks (Class Pelecypoda = Bivalvia) [173] and other classes of mollusks [174]. In these studies, he classified the types of sexuality by first establishing two main groups: (I) ambisexuality (monoecism, hermaphroditism), when the gonad is capable of expressing both sexual phases (male and female) in the lifetime of the organism; and (II) unisexuality (dioecism, gonochorism), when the gonad is capable of expressing only one sexual phase, male or female, in the lifetime of the organism. He also established four categories of ambisexuality: (1) functional ambisexuality (functional hermaphroditism), when the gonad expresses both sexual phases simultaneously and both phases are functional (i.e. both phases produce functional gametes); (2) consecutive sexuality, when the gonad experiences a sex change only once in the lifetime of the organism, usually from male to female; (3) rhythmical consecutive sexuality, when the gonad undergoes several sex changes (two or more) in the same reproductive season, as in the larviparous oysters (Ostrea edulis, Ostrea lurida); (4) alternative sexuality, when the gonad undergoes a sex change – but not always – from one reproductive season to the next, as in oviparous oysters (Ostrea virginica, Crassostrea gigas). All these categories of sexuality, as well as the sex change, are described and defined upon the concept of germline sex and not upon the concept of somatic sex, which remains to be studied and defined in bivalve mollusks. As a result of our laboratory research, histological images of gonads in several species of bivalve mollusks exhibiting different types of sexuality as well as sex change are shown and explained: Mitylus galloprovincialis (Fig. 11.1 (a, b)); Pecten maximus (Fig. 11.1 (c), Fig. 11.2 (a)); Aequipecten opercularis (Fig. 11.2 (b, c)); Chlamys varia (Fig. 11.3 (a–c); Crassostrea gigas (Fig. 11.4 (a–c), 11.5 (a–c)); Ostrea edulis (Fig. 11.6 (a–c), Fig. 11.7 (a–c)). Mytilus galloprovincialis (Fig. 11.1 (a, b)) is a dioecious species of Mytilidae family whose members only develop a germline sex (male or female) in their lifetime and do not change sex. The male form is shown in Fig. 11.1 (a); within a follicle (F),

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the germline component of the gonad displaying the germline sex consisting of spermatogonia, the process of spermatogenesis and spermatozoa can be seen; the follicle wall (Fw) is marked consisting of a thin line of fibroblasts. In our view, this follicular wall could constitute the somatic component of the gonad which might in some way express the male somatic sex of this animal. The female form is presented in Fig. 11.1 (b); within a follicle (F) the germline component of the gonad displaying the germline sex consisting of several ripe oocytes is shown; it is marked as the follicle wall (Fw) consisting of a thin layer of fibroblasts that could constitute the somatic component of the gonad which might somehow express the female somatic sex in this mollusk. Fw

ADG Fw

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Fig. 11.1: Trichrome stain. (a) Male gonad of Mytilus galloprovincialis. A gonadal follicle (F) with the male germline component exhibiting the male sexual phase (MSP) in spermatogenesis and with spermatozoa (Spz) is shown; the follicle wall (Fw) is marked; ADG: adipogranular cells. (b) Female gonad of Mytilus galloprovincialis. A gonadal follicle (F) is displayed with the germline component showing female sexual phase consisting of ripe oocytes (Ooc); the follicle wall (Fw) is marked; ADG: adipogranular cells. (c) Gonad of Pecten maximus. The simultaneous development of both male and female sexual phases of an individual (MG: male gonad; FG: female gonad) is shown; the male sexual phase is represented by follicles containing sperm; the female sexual phase consists of follicles containing oocytes (Ooc) at different stages of development; follicular walls (Fw) are marked.

Pecten maximus (Fig. 11.1 (c), Fig. 11.2 (a)) and Aequipecten opercularis (Fig. 11.2 (b, c)) are ambisexual species of the Pectinidae family whose members show functional ambisexuality (functional hermaphroditism). Individuals of these species develop both male and female sexual phases simultaneously to maturity and do not change sex. Each sexual phase develops in its own follicles. Figures show male and female follicles developing inside the sexual phase. Follicular walls are marked (Fw); they consist of a thin layer of fibroblasts constituting the gonad somatic component that could express the somatic sex. Along the border between the male and female sexual phases, the follicular walls of male follicles and those of female follicles run parallel and very close to each other. In these mollusks, we have the case where the gonad consists of two germline components (male and female) developing simultaneously, as well as two somatic components: the fibroblastic follicular walls of both male and female follicles. Whether the somatic component of each sexual phase expresses its own somatic sex (male or female) and the way how this component influences the expression of each germline sex in these species are issues that remain to be studied.

11 Climate change and reproduction, sex and sex change in shellfish

Fw FG

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Fig. 11.2: Trichrome stain. (a) Gonad of Pecten maximus. The simultaneous development of both male and female sexual phases (MG: male gonad; FG: female gonad) of another individual is shown; the male sexual phase is represented by follicles containing spermatozoa, the female sexual phase consists of follicles containing oocytes (Ooc) at different stages of development; follicular walls (Fw) are marked. (b–c) Gonads of Aequipecten Opercularis. The gonad of two individuals with both sexual phases male and female developing simultaneously (MG: male gonad; FG: female gonad) is displayed. The male sexual phase is represented by follicles in spermatogenesis and with spermatozoa, and the female sexual phase consists of follicles containing ripe oocytes (Ooc); follicular walls (Fw) are marked.

Chlamys varia (Fig. 11.3 (a–c)) is an ambisexual species of Pectinidae family whose members display consecutive sexuality; individuals of this species change sex only once in their lifetimes, and the sex change usually occurs from male to female. Individuals of this species develop germline sex as male (Fig. 11.3 (a)) or female (Fig. 11.3 (b)) and sometimes some individuals change sex from male to female (Fig. 11.3 (c)); when a sex change occurs, both male and female sexual phases develop together inside the same follicle which is known as the intersexual follicle. Fig. 11.3 (a) shows several gonadal follicles of a male individual developing the male sexual phase to maturity; the follicles are filled with sperm. Once again the gonad somatic component is indicated as the follicular wall consisting of a fibroblastic thin layer. Fig. 11.3 (b) shows several follicles of the gonad of a female individual where the germline component displays the germline sex as oocytes at different stages of development (vitellogenic and postvitellogenic oocytes); the somatic component is marked as the follicular wall constituted by fibroblasts. Fig. 11.3 (c) shows the gonad of an individual changing sex from male to female; both male and female sexual phases develop within the same follicle (intersexual follicle), the male sexual phase being more advanced in maturation (spermatozoa) than the female sexual phase (previtellogenic and vitellogenic oocytes). The onset of sex change occurs during an annual reproductive cycle just before, at or after the male sexual phase, which is more advanced in maturation, and causes the spawning of gametes. The somatic gonad component is shown as the follicular wall constituted by a thin layer of fibroblasts. It is not known what might be the endogenous signaling processes and molecular mechanisms causing the onset of the female sexual phase within follicles that previously developed the male sexual phase which results in a sex change. Crassostrea gigas (Fig. 11.4 (a–c), Fig. 11.5 (a–c)) is an ambisexual oviparous species of the Ostreidae family whose members display alternative sexuality; adult

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Fig. 11.3: Trichrome stain. Gonads of three individuals of Chlamys varia. (a) Gonad of a male individual. The germline component is represented by several follicles (F) containing spermatozoa (Spz); follicular walls (Fw) are indicated. (b) Gonad of a female individual. The germline component is represented by several follicles (F) containing vitellogenic (vOoc) and ripe (mOoc) oocytes; follicular walls (Fw) are indicated. (c) Gonad of an individual changing sex from male to female. Both male and female sexual phases develop inside the same follicle. Male sexual phase locates in the lumen of follicle, is more advanced in development than female phase and is represented by spermatozoa (Spz); female sexual phase is leaning against the follicular wall (Fw) and is represented by previtellogenic (pOoc) and vitellogenic (vOoc) oocytes.

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Fig. 11.4: Hematoxylin-eosin stain. Gonads of three individuals of Crassostrea gigas. (a) Gonad of a male individual. The germline component is represented by two follicles (F) containing spermatozoa; follicular walls (Fw) are marked. (b) Gonad of a female individual. The germline component is represented by several follicles (F) containing ripe oocytes (mOoc); follicular walls are marked. (c) Gonad of an individual changing sex from male to female. Both male and female sexual phases develop inside the same follicle (F). The male sexual phase is more advanced in development than the female phase and is represented by primary spermatocytes at various stages of Meiosis-I (MSP-Me); female sexual phase leaning against the follicular wall (Fw) is represented by oogonia in Meiosis (zygotene/pachytene stages of Meiosis-I) (FSP-Me). VC: vesicular cell.

individuals develop only a sexual phase (male or female) each reproductive season and usually change sex alternately, but not always, from a reproductive cycle to the next; they normally start with a change from male to female. Sex change occurs when the animal spawns the current sexual phase that develops to maturity. An individual may experience sex changes from male-female or female-male alternately; both sexual phases develop together inside the same follicle (intersexual follicle) during the onset of sex change. Fig. 11.4 (a) shows two gonadal follicles of a male individual; the gonad germline component is represented by sperm, and the marked follicular walls, constituted by fibroblasts, are the gonad somatic component. Fig. 11.4 (b) illustrates several gonadal follicles of a female individual where the gonad germline component consists of ripe oocytes (postvitellogenic oocytes), and the gonad somatic component

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is represented by the follicle fibroblastic walls. Fig. 11.4 (c) and Fig. 11.5 (a) represent the gonad of two individuals changing sex from male to female. In Fig. 11.4 (c), the male sexual phase (MSP) is represented by the process of spermatogenesis with primary spermatocytes at various stages of Meiosis-I, and the female sexual phase (FSP) is represented by the process of oogenesis with oogonia at the zygotene/pachytene stages of Meiosis-I. FSP is leaning against the follicular wall; MSP is more advanced than FSP in the gametogenic process. In Fig. 11.5 (a), MSP is represented by sperm in the follicle lumen, and FSP is represented by the process of oogenesis with oogonia at the zygotene/pachytene stages of Meiosis-I; again, FSP is leaning against the follicular wall. Fig. 11.5 (b, c) shows the gonad of two individuals that changed sex from female to male; in both cases MSP is the sexual phase developing in the current reproductive cycle and is represented by sperm, and FSP is represented by a few non-spawned ripe oocytes remaining from the previous reproductive cycle in which the individual behaved as a female.

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Fig. 11.5: Hematoxylin-eosin stain. Gonads of three individuals of Crassostrea gigas changing sex. (a) Gonad of an individual changing sex from male to female. Male sexual phase is located in the follicle lumen, is more advanced in development than the female phase and is represented by spermatozoa (Spz); the female sexual phase leaning against the follicular wall (Fw) is represented by oogonia in Meiosis (zygotene/pachytene stages of meiosis-I) (FSP-Me). (b) Gonad of an individual that changed sex from female to male. Both male and female sexual phases are located in the genital channel (GC). The male sexual phase (MSP) belongs to the current reproductive cycle and is represented by spermatids and spermatozoa; the female sexual phase belongs to the previous reproductive cycle and is represented by one non-spawned ripe oocyte (Ooc). (c) Gonad of an individual that changed sex from female to male. Both male and female sexual phases develop inside the same follicle (F). The male sexual phase (MSP) belongs to the current reproductive cycle and is represented by spermatozoa; the female sexual phase belongs to the previous reproductive cycle and is represented by two non-spawned ripe oocytes (Ooc).

Ostrea edulis (Fig. 11.6 (a–c), Fig. 11.7 (a–c)) is an ambisexual larviparous species of the Ostreidae family whose members display rhythmical consecutive sexuality, where the gonad undergoes several sex changes (two or more) in the same breeding season. These mollusks develop both male and female sexual phases simultaneously within the same follicle (intersexual follicle), but a sexual phase is more advanced in development and will ripen before the other; the more ripe sexual phase is the phase to be discharged at the next spawning. This phase, known as functional sexual phase, gives

390 | Joaquín Espinosa, Sara Silva-Salvado, and Óscar García-Martín

F Sb

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Fig. 11.6: Trichrome stain. Gonads of three individuals of Ostrea edulis. (a) Gonad of an individual behaving as a functional male. The functional male sexual phase located inside the genital channel (GC) is shown and represented by spermatozoa grouped in sperm balls. (b) Gonad of an individual behaving as a functional female. The functional female sexual phase inside two follicles (F) is shown and represented by ripe oocytes (mOoc); follicular walls (Fw) are marked. (c) Gonad of an individual developing both male and female sexual phases simultaneously; the male sexual phase is more advanced in development than the female sexual phase and is represented by spermatogenesis and primary spermatocytes (Spc); the female sexual phase is represented by oocytes at different stages of ripening (Ooc).

Ooc F

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Fig. 11.7: Trichrome stain. Gonads of three individuals of Ostrea edulis. (a) Gonad of an individual developing both sexual phases, male and female, simultaneously inside the same follicle (F). The male sexual phase, represented by spermatogenesis and grouped spermatozoa (Spz), is more advanced in development than the female sexual phase and is the functional sexual phase and the phase to be spawned first. Consequently, the individual will behave as a functional male in the next spawning. The female sexual phase is represented by vitellogenic oocytes (Ooc) and this phase will be subsequently spawned, then the individual will behave as functional female; follicular walls (Fw) are marked. (b) Gonad of an individual that completed the current reproductive cycle as a functional male and currently is at the resting stage. Non-spawned sperm balls (Sb) are shown inside a follicle (F); this individual will probably start the next reproductive cycle as a functional female; the follicular wall is marked (Fw). (c) Gonad of an individual that completed the current reproductive cycle as a functional female and is currently at the resting stage; several empty follicles (F) and two follicles containing non-spawned ripe oocytes (Ooc) can be observed; this individual will probably start the next reproductive cycle as a functional male.

the current functional sex to the animal; thus the animal behaves with a functional sex (male or female) at spawning and with the opposite functional sex at the next. Fig. 11.6 (a) shows an individual behaving as a functional male where MSP is developed to maturity and is represented by spermatozoa grouped in sperm balls within the genital channel, a characteristic feature of this species. Fig. 11.6 (b) displays an animal behaving functionally as female, where FSP is represented by ripe oocytes;

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the follicular wall is marked as the gonad somatic component. Fig. 11.6 (c) shows an animal with both sexual phases MSP and FSP developing simultaneously within the same follicles, and where MSP is more advanced in development than FSP. Fig. 11.7 (a) displays the gonad of an individual developing both sexual phases MSP and FSP; MSP (spermatogenesis and spermatozoa) is more advanced in development than FSP (vitellogenic oocytes); therefore, MSP will be discharged at the just next spawning and the current functional sex of the animal is male. At the following spawning, the functional sex of the animal will be female and the mollusk will behave as a female. Fig. 11.7 (b) displays the gonad of a male individual (currently behaving as functional male) that has completed the spawning of the male sexual phase; the follicles are empty and some remaining non-spawned sperm balls can be observed inside. At this point in time, the gonad is at the resting/restoration stages of the current reproductive cycle, and the animal will probably start the next reproductive cycle as a functional female. Fig. 11.7 (c) shows the gonad of a female individual, currently behaving as functional female, that has completed the spawning of the female sexual phase. The follicles are empty and a few remaining non-spawned ripe oocytes can be seen inside. At this point in time the gonad is at the resting/restoration stages of the current reproductive cycle, and the animal will probably start the next reproductive cycle as a functional male. With regard to the sex change, mollusks of the species Ostrea edulis behave differently than those of the Chlamys varia and Crassostrea gigas species. While the latter change sex from one reproductive cycle to the next (i.e. between reproductive seasons), the former really do not change sex; instead, they relieve one sex by the opposite one or more times within the same reproductive season. Ostrea edulis develops both sexual phases male and female simultaneously and delayed to each other within the same reproductive season, ripening and spawning a sexual phase one time and the opposite the next.

11.3.7 What does sex change mean and how could this process be performed by bivalve mollusks? An adult organism changes sex in its sexual life history when its gonad changes from producing a gamete phenotype (spermatozoon or oocyte) to developing the other; i.e. when a change in its gonad germline sex is observed. In such organisms the change of sex can occur only once or several times in their lifetime. Except in some fish species [172], sex change in adult organisms is infrequent in vertebrates where most species are dioecious. In light of current knowledge, and as mentioned above, when a vertebrate animal is born, it comes equipped from embryonic life with a sexually differentiated gonad constituted by its somatic and germline components, and it is therefore endowed with both a somatic sex – expressed in the gonad somatic component – and a germline sex – expressed in the gonad germline component. These two types of sex are consistent with each other; these animals usually do not change sex. At present,

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it is not known how bivalve mollusks determine the somatic sex in the embryo, nor is it known whether GSD or TSD mechanisms might be involved in this process. It is also not known how these animals – also in early embryonic life – differentiate the sex of the primordial germ cells (PGC) to become sexualized germinal stem cells (GSC). However, based on the expression of germline sex in adult individuals, it is known that there are both dioecious and ambisexual species in bivalve mollusks; it is also known that individuals of ambisexual species are sex changers. What would have to happen for a bivalve mollusk to change sex? In our view, and in the light of current knowledge, three conditions have to be met: (1) individuals have to be ambisexual; (2) in the adult animal, the gonad somatic component should hold the sexual bipotentiality that the embryonic genital ridge had and where such a component originates; and (3) in the adult, GSC cells should retain the sexual bipotentiality that embryonic PGC cells had and from which GSC cells derived. Conditions (2) and (3) would enable the mollusk to express both male and female sexual phases in its lifetime, and also would enable the individual to express such sexual phases in different ways: simultaneously, consecutively, alternatively or rhythmically. As it is currently known in vertebrates, the gonad germline sex is induced in a coherent way by the gonad somatic sex expressed by the gonad somatic component. Then, for a mollusk to change sex, it should change the sex of its gonad somatic sex; therefore the sex change would consist of a gonad somatic sex change which, in turn, would consistently induce a gonad germline sex change. This hypothesis is based on both the sexual duality and flexibility of the gonad somatic sex and the sexual bipotentiality of the GSC cells as well. Then, for a mollusk to change sex, it has to accomplish two tasks: (1) to turn off the somatic sex (male or female) that the gonad somatic component is currently expressing; and (2) to turn on the opposite somatic sex that the gonad somatic component is also able to express. Sufficient and profound research issues arise from this approach. (1) In adult individuals of bivalve mollusks, whether they change sex or not, where is the gonad somatic component, i.e. what cell types constitute it? Nothing is known about this issue, and in this regard we propose that the gonad somatic component could consist of fibroblasts forming the wall of gonadal follicles; furthermore, little information exists on the non-germline cell types existing inside the gonadal follicles and their function [175]. (2) If so, how does the somatic component express somatic sex? (3) How does the somatic sex influence the GSC cells to differentiate the germline sex? All of these processes might involve both networks of genes and cell signaling pathways where certain types of proteins should be involved in the processes of activation/deactivation of genes. These protein-gene interactions could be affected by temperature, and therefore temperature changes caused by climate change could affect sex expression in both dioecious species of bivalve mollusks and sex change in ambisexual species of bivalve mollusks whose individuals are sex changers.

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11.3.8 Temperature, photoperiod, reproductive cycle and sex change in bivalve mollusks The knowledge of the reproductive cycle of a biological species is essential for understanding its life history and to manage any decision successfully that, from a human perspective, is to be made about it. The concept of the reproductive cycle of bivalve mollusks has already been defined in Section 11.3.1, and it is an update of what was stated by Seed [181], who, in 1976, defined the reproductive cycle of a bivalve mollusks as the overall processes occurring cyclically, starting from the gonad activation and going through gametogenesis, spawning and subsequent gonadal recession. The reproductive system of bivalve mollusks is generally anatomically quite simple and consists of primary genital channels (gonoducts) and numerous smaller channels that lead to a large network of follicles [182]. Generally, the gametes are expelled through own openings independent of the renal system; but in the case of pectinid species (scallops), such release occurs through the renal channels [183]. The reproductive cycle of bivalve mollusks has been described in a large number of studies, by a large number of authors and in very different parts of the world [184–187]; the references quoted here wish to be an act of sincere and deserved recognition by the authors of this work to two pioneers and essential researchers in this area of knowledge, people who, moreover, we have known for a long time. The sequence of stages of the reproductive cycle can be synchronous among individuals in a population or asynchronous in the event that groups of individuals spawned at different times of the year. Barber and Blake [188] noted that in scallop populations, individuals tend to mature and spawn synchronously, but variations can occur in the frequency, timing and duration of the spawning, depending on environmental variations between years or even between different locations. Bivalve mollusks have the ability to store nutrients (glycogen, lipids and protein) into specialized tissues (muscle, digestive gland and, in some cases, the mantle) to meet the metabolic demands of gametogenesis. In addition to these tissues, marine bivalve mollusks also have different types of somatic cells that store and mobilize nutrients, such as vesicular cells (VC) – glycogen storage containers – and in some groups such as mussels adipogranular cells (ADG) that store proteins, lipids and glycogen [189, 190]. Bivalve mollusks display a seasonal cycle of storage and mobilization of energetic reserves, which correlates with the annual reproductive cycle and food availability [184, 188]. This cycle of storage and use of reserves is closely related to the reproductive cycle and is a regulatory factor both in reproduction and gamete production. Gabbott [191] noted that the use of energy reserves depends on the food availability in the environment, the stage of gonadal development and the general metabolic activity of the mollusk species. Bayne [187] and Kang et al. [192] classified bivalve mollusks based on their reproductive strategy. This strategy may be: (1) conservative: one in which the seasonal stor-

394 | Joaquín Espinosa, Sara Silva-Salvado, and Óscar García-Martín

age cycle of energy reserves is clearly separated from the reproductive cycle; (2) opportunistic: one in which the storage cycle of energy reserves and the reproductive cycle occur simultaneously; and (3) mixed: an intermediate strategy between conservative and opportunist. Sastry [182] noted that in bivalve mollusks, the time and duration of the reproductive cycle can be considered as determined by the interaction between various endogenous and exogenous factors. For this author, the reproductive cycle is a response, genetically controlled, to environmental conditions. Among the many factors that, according to the literature, may regulate the reproductive cycle of bivalve mollusks [193], the following can be highlighted:

(1) Regulation by exogenous factors (1a) Latitude From a bio-geographical point of view, it is accepted, in general, that a pattern exists in which the spawning periods of bivalve mollusks species are broader the more they are distributed in Southern latitudes. This variation of the reproductive period as a function of latitude may be due to the combined effect of change in latitude and temperature; since both factors vary simultaneously, it is difficult to identify the effect of each one. Furthermore, the spawning period of bivalve mollusks tends to be seasonally later in the populations located further to the North [184].

(1b) Temperature The temperature, in addition to having great importance in the geographical distribution of species, significantly influences the time of spawning. The influence of temperature on the reproduction of marine invertebrates has been widely analyzed for almost 100 years [194–197]. It is widely accepted that boreoartic species have a critical temperature above which they are unable to perform spawning, while species of temperate and tropical waters have a critical temperature below which are unable to produce larvae [198]. Variations in seawater temperature seem to induce spawning in many species, although as the season progresses, this synchronization seems lost in many of them [199]. The process more directly subject to regulation by water temperature is the gametogenic cycle [184, 200]. Once gametogenesis is started, its development varies depending on water temperature. This parameter acts as synchronizer for the seasonal rhythms of that process. Lucas [199] and Burnell [201] reported that in France and Ireland the minimum temperature for the variegated scallop (Chlamys varia) to spawn is 15 °C; however, exceptions have been demonstrated. Since this pectinid has an opportunistic reproductive strategy, if environmental conditions are favorable for early gonad development, spawning can be accomplished even without reaching that minimum temperature.

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Temperature variations of only 1–2 °C, increasing as well as decreasing, can stimulate the spawning of different bivalve mollusks species [201, 202]; in this way, the number of spawnings may vary annually according to the local temperature regime existing in the region where the population of individuals is located.

(1c) Abundance and availability of food The relationship between food availability in the environment and the reproductive cycle in bivalve mollusks has been widely described in the literature [186, 195, 203– 205]; the amount of food available seems to be a key factor controlling the start and end of the reproductive cycle of many species of these animals [184, 200]. Numerous cases of synchronism between massive phytoplankton blooms and the spawnings of bivalve mollusks have also been described. Various authors [202, 206] reported that the gametogenesis of different bivalve mollusks species depends on the amount of food available, matching the spawnings with peaks of chlorophyll a abundance. It has also been found that interannual differences in the temporality of the reproductive cycle have been associated with differences in the amount of available food [207].

(1d) Photoperiod Photoperiod is an environmental factor capable of regulating the reproductive cycle of bivalve mollusks, because the daily relationship between light and dark periods varies homogeneously and seasonally from year to year. It has been shown that certain species of bivalve mollusks, such as pectinid, develop gametogenesis in times with twelve hours of daylight [193]. In the northern hemisphere, as winter progresses and until the end of spring, the number of daylight hours increases, and this phenomenon coincides with the start and development of the annual reproductive cycle of bivalve mollusks. In late summer, daylight hours decrease gradually, which is accompanied by a decline in spawning activity [206]. The relevance of this environmental factor on the reproductive cycle of the pectinid Chlamys varia was demonstrated in experimental studies using different photoperiod regimes resulting in different patterns of gonadal development [208].

(1e) Salinity The influence of salinity may be important in the process of gametogenesis of bivalve mollusks (particularly in the family Pectinidae) [193], especially in areas where very sharp variations of this parameter occur. Large fluctuations in salinity may significantly affect the reproductive cycle and spawning times, but it has also been noted that, in general, bivalve mollusks (especially those that are cultivated) usually adapt to changes in salinity if such changes do not reach the lethal limits for each species [209].

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There are many other environmental factors (physical, chemical or biological) that interact by controlling, to a greater or lesser extent, the reproductive cycle of bivalve mollusks. These factors, as indicated by Barber and Blake [188], include: the lunar cycle and tides cycle; upwelling conditions; the action of waves or mechanical shocks; the depth where populations are located; the density of individuals in a population; the parasites and the diseases they may cause; the substrate nature onto which many sessile bivalve species set; the presence of contaminants in the environment; and the processes of competition between individuals in a population.

(2) Regulation by endogenous factors In addition to the above-mentioned exogenous factors that regulate a bivalve mollusk’s optimal conditions to breed in the appropriate season, the time of gonad ripening must also be controlled by an endogenous mechanism [182, 206]. This temporal endogenous regulation mechanism probably has two components: a mechanism regulating the gonadal development cycle; and a mechanism to synchronize the gonadal cycle with certain key environmental parameters. It is reported that some bivalve mollusks species may have neuroendocrine control of gametogenesis, and that cerebral and visceral ganglia of such species have neurosecretory cells displaying a neurosecretory cycle, with the appearance of neurosecretion products before the start of gametogenesis and the maximum concentration of such molecules coinciding with gametic maturation [210]. Such endogenous regulation also depends largely on the reproductive cycle pattern of bivalve mollusk species, and new neurosecretion products may form in each new gametogenesis. The existence of an annual breeding rhythm has a fundamental role in ensuring that the release of gametes occurs when environmental conditions are favorable for rapid and correct development of larvae in order to get the greatest number of surviving descendants [193]. The conclusion to be drawn from the reality just described, and this is reinforced by the literature consulted on the matter, is that the two main environmental parameters that, predictably, may influence the reproductive cycle of bivalve mollusks are temperature and photoperiod.

(3) Results from our research team To study the effects of temperature and photoperiod on different stages of the reproductive cycle in bivalve mollusks, as well as on the development of male and female sexual phases, our research team conducted a series of studies on the bivalve mollusk Ostrea edulis over the last five years [211]. The experimental design conducted made it possible to uncouple and analyze individually the effects of temperature and photoperiod on such subjects. The main results achieved are that temperature and photoperiod have: (a) direct effects on different stages of the reproductive cycle (de-

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fined in Section 11.3.1); (b) indirect effects on both the functional sexual phases and the current functional sex (both concepts defined in Section 11.3.6). The results related to each of these two parts is described below:

(A) Direct effects on the reproductive cycle stages (1) While all other factors are equal (individuals of the same origin, age, diet and other environmental variables), a short photoperiod (eight hours light per day or less) directly affects the early development of both male and female sexual phases because it slows down the development of new gametogenic series (male and female), the processes of mitosis and meiosis and vitellogenesis. This photoperiod inhibitory effect is inversely proportional to the hours of light, i.e. the fewer hours of light, the greater inhibitory effect. However, this effect is not exponential, because at zero hours of light gametogenic development is not completely stopped. (2) A long photoperiod (16 hours of light per day or more) directly affects the early development of both sexual phases, male and female, having an enhancing effect on the processes of mitosis, meiosis and vitellogenesis. This accelerator effect of a long photoperiod is directly proportional to the hours of light, i.e. more hours of light cause greater development of the early part of both sexual phases. Also, the enhancing effect of the photoperiod on the processes of mitosis favors the development of new gametogenic series. (3) Low temperature (14 °C or less) has a direct effect on the development of the reproductive cycle, slowing the progression of the second half of gametogenesis development (gamete ripening and spawning) in both sexual phases. In the case of the female sexual phase, low temperature acts as a limiting factor that inhibits spawning. Low temperature is also shown as a possible trigger factor of oocytes atresia when ripe postvitellogenic oocytes remain in the gonad for a long period. This study confirms previous results from Orton [212], who found that, for Ostrea edulis, temperatures below 16 °C inhibited the female spawning, regardless of the maturity state of the female sexual phase. (4) High temperature (18 °C or more) has a direct and accelerating effect on the second half of development (gamete ripening and spawning) of both sexual phases, male and female. The upper temperature limit to have a possible inhibitory effect on the development of both sexual phases is not clear, and such an inhibitory effect seems to be more related to the abundance or scarcity of food availability than with an elevated temperature.

(B) Indirect effects on both the functional sexual phases and current functional sex (1) Both temperature and photoperiod do not seem to have a direct effect on the type of sex, male or female, of bivalve mollusks; as a result of these studies, no signif-

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icant deviation towards one particular sex was noted. Instead, only a greater or lower development of both sexual phases, male and female, were observed. (2) However, in Ostrea edulis the expression of the individual current sex in natural environment can be modified by accelerating the relative development of one of the two sexual phases, male or female, and delaying the development and preventing the spawning of the other. Consider the following case study:

Case study In the Northern Hemisphere, an oyster population born and settled to the substrate in the month of July of a given year, starts gonad development and gametogenesis in the next autumn (from September to November). In this first year of life, nine out of ten oysters will start the reproductive cycle developing the male sexual phase, and subsequently the female sexual phase develops in the same follicles (protandry). The “normal” development – what usually happens in the natural environment – is that the male sexual phase develops first (its development also requires less time), matures and performs the spawning; therefore, at the time, most oysters have a functional male sex. Subsequent to this spawning, the female sexual phase, already incipient, develops, matures (it is already next spring) and the oysters spawn as females (although a spawning with particular characteristics for being the first spawning as females); i.e. it will have occurred as described in Section 11.3.6 and highlighting the rhythmical consecutive sexuality of this species. However, if oysters are in the early development of both sexual phases (sexual male always more developed), and the temperature and photoperiod are modified by subjecting oysters to low temperature and a short photoperiod, the female sexual phase slows its development such that, once the male phase spawned, the still immature female sexual phase does not develop and consequently does not spawn, thereby allowing the development of a new male sexual phase that develops, matures and spawns before the female sexual phase does so; therefore, the rhythmical consecutive sexuality that characterizes this species will have changed. Taken together, these results clearly manifest that temperature and photoperiod can independently affect the gametogenesis of both sexual phases, male and female, as well as the current functional sex of these mollusks, regardless of other factors such as food availability or the age of individuals that might also do so.

11.3.9 Climate change, reproductive cycle, sex expression and sex change in bivalve mollusks It’s been 40 years since Broecker [213] wondered if we were on the brink of climate change. After all that has been shown in this study, it is now clear that the answer to this question is positive. It could be debated whether, among the multiple climatic

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variables existing, some are changing more than others, that their change is not uniform in all latitudes and longitudes of the world, or whether the causes of these changes are mostly of anthropogenic origin or not; but what can not be debated is the reality of climate change and its global character. However, by systematically reviewing the numerous annually published studies reporting changes of certain physical, chemical and biological parameters, it is observed that they lose sight of some aspects of global climate change that, in our opinion, should have significant importance and be more in the forefront of the discussion in terms of their potential effects on the adaptability of different populations of bivalve mollusks inhabiting different oceans to global climate change. We are referring, firstly, to the fact that cooperative interactions between simultaneous changes of many different parameters have the ability to potentiate the effects of global climate change. This cooperation may be of a positive and reinforcing nature (synergistic effects of changes in two or more variables) or negative and debilitating (antagonistic effects). These multiple concurrent effects, which will be common in the future scenario of global climate change, will constitute a potent challenge that all forms of life on earth will face and, predictably, not all will overcome. The reality with which global climate change confronts a bivalve mollusk that inhabits the intertidal zone of any ocean is a good example of what is meant. First, it is a common situation, given the ubiquity of bivalve mollusks. It is an example that can be found on either coast of the planet. On rocks, the mollusks depend on the geographical area where they are located and the tidal cycle; they will face two different realities in the course of a day; for a part of their lives they will be covered by seawater and during the other part they won’t (the time not exposed to seawater depends on where the rocks are located). Given the predictions of global climate change, in both situations they will face a rise in the temperature of their environment, both marine and aerial. When a sessile bivalve mollusk is exposed on the rock at low tide, the organism closes its valves and retains a certain amount of seawater (intervalvar water) so it can get oxygen from it for some (limited) time, and intervalvar water is also used as a compartment to store metabolic waste products until the next high tide [214, 215]. This intervalvar water will also have a certain oxygen concentration that, as temperature increases, will be lower. When closing the valves and having a lower oxygen supply, the mollusk reduces its metabolic activity and changes its metabolism from the aerobic mode to another essentially anaerobic mode, and the individual maintains its metabolism at the expense of accumulated energetic reserves (mainly glycogen, which is a major fuel for anaerobic metabolism) [216, 217]. The metabolism, mainly anaerobic, produces progressive acidification of the internal environment of the organism [216, 217], to which an intervalvar seawater already more acidic due to the effects of global climate change must be added, as outlined above (Section 11.2.3.3).

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The seawater available to the mollusk will have, predictably, lower salinity since this is the climate change trend; also, rainfall, probably stormy, may be more frequent. Therefore, for the mollusk to keep a determined internal osmotic concentration against more dilute seawater, it will have to expend more energy [218]. As an organism equipped with a shell of calcium carbonate (calcifier organism), the mollusk will face increased energy expenditure for shell forming, because sea water will be more acidic and its content in cabonate ions (CO2− 3 ) will be lower. In addition, the individual will have to renew its shell more often, because the increased acidity of seawater and the use of the shell to buffer the body internal environment [217]; both factors will cause increased shell wear. The mollusk will have to cover all these energy needs – all the greater the more pronounced the effects of global climate change are – with energetic reserves it will have formed from filter feeding. It is likely that there will have been significant changes in the organisms that are food to a marine filter feeder mollusk (phytoplankton), changes related to both quality (size, toxin content) and quantity of phytoplankton. In any case, as the mollusk feeds by filtration, it will continue concentrating the particulate matter present in seawater in its body, and this fact will allow bivalve mollusks to be used as biomarkers to indicate, permanently, the variability that the primary productivity of the oceans will go through as global climate change progresses in this basic scenario. Therefore, bivalve mollusks may be useful biomarkers for certain aspects of climate change. In warmer and eutrophicated oceans, forecasts (Section 11.2.3) are that HAB episodes will be more frequent and even phytoplankton organisms could produce new and unknown types of toxins that eventually may be transmitted by filter-feeding organisms (bivalve mollusks) as vectors. Without excluding that this ecosystem – the origin of the oceanic primary productivity – should be controlled in the most direct way possible, the availability of organisms serving as indicators of its continuous state at all times and in different parts of the planet is an advantage to be used in the future; bivalve mollusks may be useful for this purpose. As has been shown, the challenge that global climate change presents to bivalve mollusks that inhabit the intertidal zone of the planet’s oceans is multivariable and with different climatic parameters interacting with each other. None of the models currently used to indicate which could be the result of this complexity serve as a predictive tool. However, it is true that bivalve mollusks have been living on this planet much longer than humans and, predictably, it is not the first time that they have faced changes in their environment similar to the current one; and it seems, according to research results, that they have done it in the past with quite remarkable success. It would therefore be unintelligent not to use bivalve mollusks as part of the reference organisms that, undoubtedly, will have to be identified to monitor the evolution of the effects of global climate change. The second aspect of global climate change, referred to at the beginning of this section, is called the “critical time element” by Somero and Hochachka [219]; i.e. the

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time it takes certain changes to occur, in this case applied by the authors to changes in the temperature of the marine organisms’ environment. In the cited authors’ words: The intensity of the stress arising from a change in temperature is basically proportional to the product of two terms: a) the magnitude of the temperature change; and b) the rapidity with which the change occurs.

We believe that the magnitude of global climate change that is currently happening on our planet is evidenced in everything that has been expressed in this study. Furthermore, the experimental data obtained on a daily basis indicate, without much doubt, that this change is taking place extremely fast, probably at a rate that is unprecedented in the history of our planet. And predictably all of it because of the anthropogenic contribution to climate change. As Somero and Hochachka [219] stated, time is critical to assess the capacity and possibilities of certain organisms to adapt or not to adapt to a change in a given environmental parameter. As these authors also noted, a fast enough temperature change can prevent even the simplest adaptation to change, since a quick migration to another area where the temperature is more appropriate is impossible for a sessile organism such as the bivalve mollusk we are considering. All changes in environmental parameters experienced by a sessile bivalve mollusk inhabitant of the intertidal zone are changes occurring in a very short time, in hours, which may be included in what is called “instantaneous changes of these parameters” [219]. To successfully respond to these changes, the organism must possess a number of immediate adaptation mechanisms that are characteristic and can be activated at the time the change occurs [220]. Bivalve mollusks have this capacity of adaptation, so they have successfully colonized these habitats. When there is enough time to adapt to change (months, years, centuries or millennia), other adaptations come into play that usually involve changes in the organisms’ genetic endowments [112]. Bivalve mollusks, along their wide life trajectory on the planet, have diversified into a wide range of species that are constantly showing an enormous plasticity and capacity to adapt to the challenges of global climate change. When studying the latest works by experienced researchers [44–46] trying to shed light on features that may favor some organisms compared to others in overcoming the challenges of global climate change (i.e. determining the organisms that will be “winners” and which will be “losers”), different species of bivalve mollusks always appear among the chosen cases. One of the methods that provides better results on the degree of necessary change necessary to overcome environmental parameters at the molecular level has been to study and compare the sequences of enzymes that perform the same function (orthologous enzymes) in different species of the same genus that share the same habitat but have a different distribution for a given environmental parameter (temperature, salinity and oxygen availability) [44–46]. Such studies have revealed that a single amino acid change in a protein of 330 amino acid residues, or re-

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placement of a single nucleotide in a gene sequence with more than 1000 nucleotides in its structure (0.1 % change in the sequence), is sufficient to produce an adaptive change in protein function [44–46]. As stated in Section 11.3.8, temperature and photoperiod directly affect the reproductive cycles of bivalve mollusks, and indirectly affect both the expression of sexual phases and sex expression as defined in Section 11.3.6. Photoperiod is one of the few environmental parameters that, so far, does not seem to be affected by the changes caused by global climate change; however, this does not apply, as has been seen, to temperature. Global warming will impact the reproductive cycle of bivalve mollusks and accelerate, in all probability, the maturing of sexual phases and increase the number of spawnings, as is already the case today with bivalve mollusks species that have a subtropical distribution. The cellular and molecular mechanisms underlying the processes of sex determination and the sexual differentiation of the gonad and the determination of primordial germ cells (PGCs) of bivalve mollusks species are currently unknown; and the possible evolutionary advantages in the presence of so many different types of sexuality and the reality of sex changes in this group of organisms has also not been explained. Given the evident evolutionary success of this animal phylum, it is proposed to use the phylum Molluska, and in particular the Bivalvia class, as “reference organisms” to deepen the understanding of the cellular and molecular basis of their biological characteristics, in the belief that these studies will provide important information for planning and managing the future we are facing with the reality of global climate change. In this sense, and reaffirming the belief of this work’s authors in August Krogh’s known principle, masterfully formulated by Sir Hans Krebs [45, 220]: For many problems there is an animal on which it can be most conveniently studied,

the authors propose, for the specific case of the effects caused by global climate change, to replace the word “animal” by “bivalve mollusk” in such a scientific and wonderful statement.

11.4 Concluding remarks As a result of this study, our conclusions and views are the following. (1) Global climate change is currently unequivocal. Various physical, chemical and biological parameters in the ocean environment are affected: temperature, salinity, CO2 concentration, pH, oxygen concentration, light, nutrients and primary production, among others. Various marine organisms such as bivalve mollusks also experience changes in their reproductive cycles and sex expression. (2) In recent years, the HAB episodes have been increasing. Even though it is not excluded that the importance of this apparent increase may have been the effects of a

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progressive ocean eutrophication due to anthropogenic origins, it is increasingly evident that global climate change may be one of the causes of such increases. (3) The species of bivalve mollusks are important links in the marine food chain, and plenty of these species are of economic importance to humans. The sex ratio is a critical parameter for the reproductive sustainability of natural and farmed populations of these species over time, and this sex ratio must be about 1 : 1 for dioecious species, for species that have functional ambisexuality and for species whose individuals are sex changers. (4) In a population of bivalve mollusks, an increase of temperature in the environment of individuals could skew the sex ratio favoring either male or female sex in the following scenarios. (I) in dioecious species: (a) during embryonic development of individuals, by favoring the expression of either male or female somatic sex; (b) during the reproductive cycle of individuals, by favoring either of the gametogenic processes: spermatogenesis or oogenesis. (II) In ambisexual species: whose individuals are sex changers, by altering the pattern of sex change, thus favoring the expression of a particular sexual phase, male or female, to the detriment of the other. (III) In species with functional ambisexuality: by altering the simultaneous expression of both male and female sexual phases, thus favoring the expression of a particular sexual phase, male or female, over the other. (IV) An increase in temperature could also influence: (a) the temporal pattern of the reproductive period of bivalve mollusks, shifting stages of the reproductive cycle seasonally backwards compared to the seasonal pattern it currently displays; (b) the duration of certain stages of the reproductive cycle, by shortening or lengthening the temporal extent of the stage. Therefore, if the trend of climate change increases the temperature of bivalve mollusk species’ marine environment, it could affect the sex ratio of this species and, consequently, their reproductive sustainability. (5) It is currently unknown whether bivalve mollusks are in accordance with current knowledge about gonad formation in the embryo, and if so, then it is unknown. (a) What are the cells constituting the somatic component of the gonad in the adult animal? In this regard, the authors hypothesize that such cells might be the fibroblasts that form the follicular walls of the gonad. (b) What are the cellular and molecular processes underlying sex determination and differentiation of embryonic gonad in both dioecious and ambisexual species? (6) A framework hypothesis and experimental studies are needed to explain the cellular and molecular basis underlying the ambisexuality and sex change of bivalve mollusks; in this regard, the authors put forward the following hypothesis. (a) The sexual bipotentiality of the embryonic genital ridge should be maintained by the somatic component of the gonad – possibly the fibroblasts forming the follicular walls – in the adult animal. (b) The sexual bipotentiality of primordial germ cells (PGC) of the embryo should be maintained by the germline stem cells (GSC) in the adult animal. (c) In the adult animal, cellular and molecular interactions between the somatic component of the gonad – the fibroblasts of follicular walls – and the GSC cells could

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determine the sexual differentiation of these cells to become spermatogonia or oogonia. This could be a way to explain the ambisexuality, functional ambisexuality and sex change in bivalve mollusks. (7) In bivalve mollusks, it should be attempted to apply the concepts of: (a) gonadal sex determination and differentiation during the embryonic life of the individual; (b) PGC cell determination and their differentiation into GSC cells during the embryonic life of the individual; (c) gonad somatic sex; and (d) germline sex. Pursuing this path, the gene and molecular mechanisms underlying these processes should be elucidated. (8) Global warming will impact the reproductive cycle of bivalve mollusks by accelerating, in all probability, the maturing of sexual phases and increasing the number of spawnings, as is already the case today with bivalve mollusks species in subtropical distributions. (9) It is proposed that bivalve mollusks, organisms with ubiquitous distribution in all the world oceans, could be “reference biomarker organisms” to study and monitor the impacts of global climate change. The task is to find and select suitable species.

Acknowledgment The authors of this study are grateful to Professor Luis M. Botana for his invitation to develop and write this chapter.

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[173] Coe W. Sexual differentiation in mollusks. I. Pelecypods. The Quarterly Review of Biology 1943;18:154–64. [174] Coe W. Sexual differentiation in mollusks. II. Gastropods, Amphineurans, Scaphopods and Cephalopods. The Quarterly Review of Biology 1944;19:85–97. [175] Franco A, Berthelin H, Goux D, Sourdaine P, Mathieu M. Fine structure of the early stages of spermatogenesis in the Pacific oyster, Crassostrea gigas (Mollusca, Bivalvia). Tissue and Cell 2008;40:251–60. [176] Warner D, Shine R. The adaptive significance of temperature-dependent sex determination in a reptile. Nature 2008;451:566–8. [177] Hulin V, Delmas V, Girondot M, Godfrey M, Guillon J. Temperature-dependent sex determination and global change: are some species at greater risk? Oecologia 2009;160:493–506. [178] Mitchell N, Janzen F. Temperature-dependent sex determination and contemporary climate change. Sex Dev 2010;4:129–40. [179] Kallimanis A. Temperature dependent sex determination and climate change. Oikos 2010; 119:197–200. [180] Pen I, Uller T, Feldmeyer B, Harts A, While G, Wapstra E. Climate-driven population divergence in sex-determining systems. Nature 2010;468:436–9. [181] Seed R. Ecology. In: Marine mussels: their ecology and physiology. Bayne BL. [ed.] Cambridge, UK; Cambridge University Press: 1976. p. 13–65. [182] Sastry AN. Pelecypode (excluding Ostreidae). In: Reproduction of marine invertebrates. Vol V. Molluscs: pelecypods and lesser classes. Giese AC and Pearse JS. [eds.] New York; Academic Press: 1979. p. 131–292. [183] Lubet P. Essai d´analyse experimentale des perturbations produites par las ablations des ganglions nerveux chez Mytilus edulis L. et Mytilus galloprovincialis Lmk (Mollusques Lamellibranches). Ann Endoc 1966;27:353–65. [184] Lubet P. Recherches sur le cycle sexuel et l´emission des gametes chez les mytilides et les pectinides. Ph D Thesis Rev Trav Off Pêche Marit 1959;23(4):387–548. [185] Lubet P, Faveris R, Besnard JV, Robbins I, Duval P. Annual reproductive cycle and recluitment of scallop Pecten maximus (Linnaeus, 1758), from the bay of Seine. In: An International Compendium of Scallop Biology and Culture. Shumway SE, Sandifor PA. [eds.] New York, USA; The World Aquaculture Society: 1991. p. 87–94. [186] Bayne BL. Reproduction in bivalve molluscs under environmental stress. In: Physiological ecology of estuarine organisms. Wiley ML. [ed.] Columbia; University South Carolina Press: 1975. p. 259–77. [187] Bayne BL. Aspects of reproduction in bivalve molluscs. In: Estuarine processes, Vol. I: Uses, stresses and adaptation to the estuary. Wiley ML. [ed.] New York; Academic Press: 1976. p. 432–48. [188] Barner BJ, Blake NJ. Reproductive physiology. In: Scallops: Biology, ecology and aquaculture. Developments in aquaculture and fisheries science Vol. 21. Shumway SE. [ed.] Amsterdam; Elsevier Science Publishers BV: 1991. p. 377–428. [189] Pipe RK. Ultrastructural and cytochemical study on interactions between nutrient storage cells and gametogenesis in the mussel Mytilus edulis. Mar Biol 1987;96(4):519–28. [190] Lenor F, Robbins I, Mathieu M, Lubet P, Gabbott PA. Isolation, characterization and glucose metabolism of glycogen cells (vesicular connective–tissue cells) from the labial palps of the marine mussel Mytilus edulis. Mar Biol 1989;101:495–501. [191] Gabbott PA. Developmental and seasonal metabolic activities in marine molluscs. In: The mollusca. Vol 2. Environmental biochemistry and physiology. Hochachka PW. [ed.] New York; Academic Press: 1983. p. 165–217.

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[192] Kang CK, Park MS, Lee PY, Choi WJ, Lee WC. Seasonal variations in condition, reproductive activity, and biochemical composition of the Pacific oyster, Crassostrea gigas (Thunberg), in suspended culture in two coastal bays of Korea. J Shellfish Res 2000;19:771–8. [193] Iglesias P. Estudio de la reproducción y reclutamiento de los pectínidos de interés comercial de las Rías Gallegas. Tesis Doctoral. Universidad de Santiago de Compostela; 2012. [194] Orton JH. Sea temperature, breeding and distribution in marine animals. J Mar Biol Ass UK 1920;12:339–66. [195] Sastry AN. Temperature effects in reproduction of the bay scallop, Aequipecten irradians Lamarck. Biol Bull 1966;130:118–34. [196] Loosanoff VL. Development of shellfish culture techniques. In: Proc. of he conference on artificial propagation of commercially valuable shellfish. Price KSJr, Maurer DL. [eds.] Newark; University of Delaware: 1971. p. 9–40. [197] Roman G, Acosta C. Cultivo de vieira en batea. II: Reproducción. In: Libro de Actas III Congreso Nac. Acuicult. Landin A, Cerviño A. [eds.] Santiago de Compostela; Consellería de Pesca, Marisqueo y Acuicultura: 1990. p. 461–466. [198] Corral ML, Henri G, Montes J, Polanco E. La acuicultura: biología, regulación, fomento, nuevas tendencias y estrategia comercial. Tomo I. Análisis del desarrollo de cultivos: medio, agua y especies. Fundación Alfonso Martín Escudero. Madrid; Mundi-Prensa Libros S.A.: 2000. [199] Lucas A. Recherche sur la sexualité del mollusques bivalves. Tesis Doctoral. Faculté des Sciences. Université de Rénnes (France); 1965. [200] Mackie GL. Bivalves. In: The mollusca. Vol 7. Reproduction. Tompa AS, Verdon KNH, van den Biggelaar JAM. [eds.] New York; Academic Press: 1984. p. 351– 418. [201] Burnell GM. Growth and reproduction of the variegated scallop Chlamys varia (L.) on the west coast of Ireland. Ph. D. Thesis University college Calway (Ireland); 1983. [202] Burnell GM. Annual variations in the spawning and settlement of the scallop Chlamys varia (L.) on the west coast of Ireland. In: An International Compendium of Scallops Biology and culture. Shumway SE, Sandifer PA. [eds.] Baton Rouge L.A., USA; The World Aquaculture Society: 1991. p. 47–59. [203] Pazos AJ. Relaciones entre composición bioquímica y ciclo gametogénico en la vieira, Pecten maximus L. Tesis Doctoral. Departamento de Bioquímica y Biología Molecular. Universidad de Santiago de Compostela; 1993. [204] Román G, Campos MJ, Acosta CP. Relationships among environment, spawning and settlement of queen scallop in the Ría de Arousa (Galicia, NW Spain). Aquacult Int 1996;4:225–36. [205] Román G, Campos MJ, Cano J, Acosta CP. The reproduction of the queen scallop Aequipecten opercularis in Galicia, NW Spain. In: Book of Abstracts 13th Int. Pectinid Workshop. 18–24 April. Coquimbo, Chile; 2001. [206] De la Roche JP. Ciclo reproductor y anatomía funcional de la gónada de Chlamys varia (Linnaeus, 1758). Tesis Doctoral. Universidad de Santiago de Compostela; 2008. [207] Shafee MS. Ecophysiological Studies on a temperate bivalve Chlamys varia (L.) from Lanveoc (Bay of Brest). Ph. D. Thesis. Université de la Bretagne Occidental–Brest (France); 1980. [208] Louro A, De la Roche JP, Sánchez JL, Silva A, Martínez P, Pérez-Parallé ML, Martínez I, Román G. The effect of photoperiod on the conditioning of black scallop, C. varia. I: gonad development. In: Book of abstracts of the 15th Int. Pectinid worshop, 20-26 April. Mooloolaba, Australia; 2005. p. 124–5. [209] Navarro JM, González CM. Physiological responses of the Chilean scallop Argopecten purpuratus to decreasing salinities. Aquaculture 1998;16(3-4):315–27. [210] Lubet P. Cycle neurosecretoire chez Chlamys varia L. et Mytilus edulis L. (Mollusques Lamellibranques). CR Acad Sci 1955;241:119–21.

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[211] Silva-Salvado S. Líneas germinales, ciclo reproductivo y sucesión de sexos en Ostrea edulis (Linnaeus, 1758). Tesis Doctoral. Universidad de Santiago de Compostela; 2015 (Dissertation pending). [212] Orton JH. Observations and experiments on sex-change in the European oyster (O. edulis): Part III: On the Fate of Unspawned Ova. Part IV: On the Change from Male to Female. Journal of the Marine Biological Association of the United Kingdom 1933;19(1):1–53. [213] Broecker WS. Climate change: are we on the brink of a pronounced global warming? Science 1975;189:460–3. [214] Newell RC. Biology of intertidal animals. London; Elek Books: 1970. [215] Newell RC. Adaptations to intertidal life. In: Adaptation to environment: Essays on the physiology of marine animals. Newell RC. [ed.] Denmark; Institute of Biology, University of Odense: 1976. p. 1–82. [216] Hochachka PW, Somero GN. Enzyme and metabolic adaptations to low oxygen. In: Adaptation to environment: Essays on the physiology of marine animals. Newell RC [ed.]. Denmark; Institute of Biology, University of Odense: 1976. p. 279–314. [217] Zwann A. The metabolism of invertebrate facultative anaerobes. Br Biochem Soc Symp 1976;41. [218] Lockwood APM. Physiological adaptation to life in estuaries. In: Adaptation to environment: Essays on the physidegy of marine animals. Newell RC. [ed.] Denmark; Institute of Biology, University of Odense: 1976. p. 315–92. [219] Somero GN, Hochachka PW. Biochemical adaptation to temperature. In: Adaptation to environment: Essays on the physiology of marine animals. Newell RC. [ed.] Denmark; Institute of Biology, University of Odense: 1976. p. 125–90. [220] Krebs H. The August Krogh Principle: “For many problems there is an animal on which it can be most conveniently studied”. J Exp Zool 1975;194:221–6.

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12 Effects on world food production and security 12.1 Introduction Climate change means a shift in the pattern of weather events, over the long term. In addition to natural variation, it implies increased average global temperature but also other effects that include trends towards stronger storm systems, increased frequency of heavy precipitation events and extended dry periods, rise in the sea-levels due to the contraction of the Greenland ice sheet and the decline in the volume of glaciers [1]. These changes may affect all components that influence food production and security: availability (sufficient quantity of food for consumption); access (ability to obtain food regularly); stability (risk of losing access to resources required to consume food); and utilization (quality and safety of food) [2]. Today already 40 % of the globally distributed animal food products are of aquatic origin [3]. The term aquatic food generally covers a heterogeneous group of aquatic organisms used as a source of nutrition by humans, including finfish and shellfish from marine or freshwater environments and from open seas as well as aquaculture. Even in small quantities, aquatic food can have a positive effect on nutritional status by providing essential amino acids particularly within the diet of developing countries. Fish accounts for 30 % of animal protein consumed in Asia, 20 % in Africa and 10 % in Latin America and the Caribbean [4]. This chapter aims to identify potential impacts of climate change on aquatic food safety and production, and discusses some expected effects that are supported by data; it also considers other issues that are speculative. The food safety issues cover agents of foodborne and waterborne disease including zoonotic diseases, environmental contaminants with significance to the food chain and biotoxins in water, fishery and aquaculture products. The chapter will dedicate specific attention to this last issue. Climate change may have both a direct and indirect impact on the occurrence of food safety hazards at any stage of the food chain from primary production through to consumption.

12.2 Foodborne and waterborne diseases People are increasingly exposed to changes in weather patterns as well as water and food quality. As climate changes are felt around the globe, these changes can impact human health in a variety of ways such as increased frequency and distribution of vector-borne disease or an increased risk of foodborne and waterborne diseases [5]. A foodborne disease is an illness caused by food, whereas a waterborne disease is attributable to drinking water. Both foodborne and waterborne diseases are important

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public problems affecting millions of people each year and resulting in substantial morbidity and mortality [6]. Some countries could find it socially and politically difficult to allocate resources to mitigate and adapt to the link between health and climate change [7]. Associations between increased incidence of food and waterborne illness and severe temperature events evidence the impact of climate change on the transmission of food and waterborne diseases [8, 9]. Global climate change could also affect food safety-inducing effects on microbial evolution and stress response, pathogen emergence and changes in water availability and quality. New demands on existing water sources occur if sea levels rise adversely; this would impact water availability [10]. Furthermore, periods of excessive precipitation and drought can influence both the availability and the microbiological quality of water. This may result in a decline in both water quantity and quality, especially in arid regions such as the Mediterranean and Northern Africa. Another emerging environmental health threat would be the drop in global freshwater resources caused by increasing rates of water extraction and contamination. Also, limited access to safe water has a negative effect on hygiene practices throughout the food chain. For instance, cholera is predominantly a waterborne disease; however, foodborne transmission can also occur through the use of contaminated water for food preparation or irrigation. There are a few characteristics of pathogens that may predispose them to being more sensitive to the impacts of global climate change. For instance, pathogens with documented stress tolerance responses (temperature, pH), such as enterohemorrhagic Escherichia coli and Salmonella compete better after adverse weather events [11]. In addition, those foodborne pathogens that cause disease at very low doses (enteric viruses, parasitic protozoa, Shigella spp., enterohemorrhagic E. coli strains) and/or have notable environmental persistence (enteric viruses and parasitic protozoa) will be of great concern.

12.3 Zoonosis and other animal diseases A Zoonotic disease is an illness that can be transmitted from animals to people in a number of ways: through direct contact with infected animals or animal products and wastes; by vectors; or through the consumption of contaminated food. On the other hand, the proliferation of zoonosis and other animal diseases may result in an increased use of veterinary drugs that could lead to possibly unacceptable levels of residues in foods [11]. Aquatic invertebrates (mollusks, crustaceans, worms, etc.) and fish are vulnerable to climate change because water is their life-support medium and their ecosystems are fragile. Certain environmental conditions are more conducive to diseases than others. For instance, warm waters can trigger disease outbreaks and cold temperatures can limit them. In this sense, significantly higher mean water temperature relative to pre-

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vious decades was thought to have contributed to a series of outbreaks of Vibrio paparahaemolyitic that occurred between 1997 and 2004 with a consequent foodborne disease attributed to the consumption of oysters harvested from Alaskan waters [11]. Similarly, an extreme rise in the fish pond water temperature has also been observed in the emergence of systemic V. vulnificus infection that occurred in Israel [12]. Temperature, salinity, and oxygen levels influence metabolic processes of fish. Those animals are poikilotherms; their internal temperature varies directly with that of their environment. This makes them very sensitive to changes in water temperature. When changes do occur, they move to areas where the external temperature is suitable. With increasing temperature, this “behavioral thermoregulation” is resulting in rapid migrations poleward or into cooler bodies of water [3]. Also, some species may already be near their upper temperature tolerances. On the other hand, a decrease in thermal movement of the water affects water quality by allowing pollutants to accumulate in upper layers of the water and has led to increased levels of mercury and other contaminants in fish. Also, reduced availability of food is even forcing aquatic species to develop longer growing periods. In addition, aquatic animals respond directly to changes in their biological environment (predators, species interactions, disease). The long-term consequences of climate change and potential environmental degradation also include aspects of disease emergence in marine plants and animals. Marine animals may be exposed to environmental stressors such as emerging or resurging pathogens and harmful algal blooms. The produced biotoxins can be transferred through the food web where they affect and even kill the higher forms of life such as shellfish, fish, birds or marine mammals and even humans that feed on them either directly or indirectly. Since many marine mammal species share the coastal environment with humans and consume the same food, they also may serve as effective sentinels for public health problems [13]. In the aquaculture sector, problems expected from a warming environment include a greater susceptibility to disease. This is particularly true for introduced aquaculture species or if engineered fish lack the innate abilities to deal with new strains of pathogens. New diseases in aquaculture could easily result in increased chemicals use. Consequently, there may be higher levels of veterinary drugs in foods [11].

12.4 Product safety in fisheries World population is expected to grow to about 9 billion by 2050. The growing need for nutritious and healthy food will increase the demand for fisheries products whose productivity is already highly stressed by excessive fishing pressure [14]. The links between fisheries and their ecosystems are deeper and more significant than those that exist in mainstream agriculture [15]. The productivity of a fishery is tied to the health and functioning of the ecosystem which it depends on for food, habitat and even seed

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dispersal. Generally the only control humans can exert over this productivity is the adjustment of fishing effort [16]. Natural climatic oscillations, particularly those at a medium (decadal) scale, have always affected fisheries (Tab. 12.1) with indirect effects [14]. They include: changes in quantity and quality of the aquatic habitat and the distribution and abundance of aquatic competitors and predators [17]; problems in food security [18]; and potential Tab. 12.1: Climate change impacts on fisheries and aquaculture: effects on aquatic food production and security [7]. Climate change

Impact on

Outcome for aquatic food

Ocean acidification

Negative effects on calciferous animals. Slowed rates of coral growth.

Declines in production.

Higher ocean temperatures

Poleward shifts in plankton and fished species. Changes in phytoplankton and zooplankton composition. Changes in physiology and sex ratios of fished species. Altered migrations and peak abundance. Increased invasive species, diseases and harmful algal blooms (HABs). Increased frequency and severity of coral bleaching events. Changes in stratification, mixing and nutrients in lakes and marine upwellings.

Changes in production and availability of fished species. Potential mismatch between prey (plankton) and predators (fished species) and declines in production. Changes in timing and levels of productivity across marine and freshwater systems. Reduced coral reef fisheries productivity. Changes in productivity and safety.

Sea level rise

Loss of coastal habitats. Saline intrusion into freshwater habitats.

Reduced production of coastal marine and freshwater systems and related fisheries.

Changes in ocean currents (ENSO and PDO)

Effects on fish recruitment. Changes in timing and latitude upwelling.

Changes in abundance of juvenile fish. Changes in pelagic fisheries distribution.

Reduced water flows and increased droughts or flods

Changes in lake water levels. Changes in dry water flows in rivers. Damage to productive assets (fish pond, rice fields).

Reduced lake and river productivity.

Increased storms frequency and severity.

Greater risk of damage in aquaculture installations.

Reduced viability of fishing and fish-farming.

Changing levels of precipitation.

Where rainfall deceases, reduced opportunities for fishing and aquaculture.

Changes in inland fishing.

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diseases [19]. Climatic direct effects on fisheries are modifications in primary production with alterations in the abundance and distribution of fish species that affect fisheries at the level of individual organisms during all life stages, species populations and communities [7, 20]. With higher water temperature, global fisheries production should remain the same; however, the spatial distribution of fish stocks may change due to the decline in primary productivity and migration of fish from one region to another in search of suitable conditions (e.g. warmer water species moving towards the poles). Those changing conditions may lead to a period of quantitative and qualitative instability in supplies. Local fishing, markets and consumers could adapt. However, instability is not favorable to international export markets [14]. Clearly, therefore, the impact of global climate change on ocean capture fisheries will be important for the availability, distribution and resilience of resources [14] as well as for food safety. An increase in water temperature also facilitates methylation of mercury and the subsequent uptake by fish; this has implications for food safety [21]. From a microbiological perspective, climate change promotes the growth of organisms such as Vibrio vulnificus; this leads to an increased risk from handling or consuming fish grown in these waters [12]. Climate interactions with species will change the composition and dynamic coupling of food webs. In some areas, reduced precipitation could lead to decreased runoff from land, starving wetlands and mangroves of nutrients damaging local fisheries. Availability of nutrients in the water affects primary productivity that also depends on ocean mixing as well as levels of light and temperature. Opposite, increased precipitation or extreme weather events, including flooding, will lead to excessive nutrient levels in rivers, lakes and coastal waters causing phytoplankton growth. Phytoplanktonic blooms are typically beneficial; however, under certain conditions, this natural phenomenon becomes harmful to other forms of life and they are named Harmful algal blooms (HABs) also known as red tides [22]. Those microalgae can also produce phycotoxins having impacts on wild (mammals, birds, fishes and invertebrates) and cultured animals (shellfish and fishes). Accumulation of phycotoxins by filter feeders (bivalve mollusks) and the subsequent consumption of these products have serious implications for humans. Cyanobacterial harmful algal blooms (CHABs) in freshwater is increasing in spatial extent and temporal frequency worldwide. Cyanobacterial blooms produce highly potent toxins and huge, noxious biomasses in surface waters used for recreation, commerce and as drinking water sources [23]. Affected water is likely to contain a variety of cyanobacterial toxins in varying concentrations that may change over the duration of the bloom. Those toxins pose a hazard to humans, domestic animals, wildlife and the ecosystem. Human exposure to the toxins can occur through ingestion of contaminated drinking water, plus dermal contact and/or inhalation of aerosols while bathing and showering in tap water or during recreational use of surface waters. Intoxication could also be due to the ingestion of contaminated fish and other foods of aquatic origin, and/or supplements. Establishing intakes and duration parameters for these ex-

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posure scenarios will facilitate the application of risk assessment approaches to these situations [24]. Looking towards the future, the question is how fisheries governance, and the national and international policy and legal frameworks, will ensure a sustainable harvest that maintains biodiversity and ecosystem functions while adapting to climate change [14]. Unfortunately, ecosystems and areas with the most potential for development of fisheries Africa, Asia and Latin America are also those most threatened by changes in climate. In those areas, fisheries could have the biggest impact on poverty reduction and food security. Many African countries are highly vulnerable to the effects on capture fisheries ecologically as well as socially and economically. This is due to their dependence on fish protein in diets, limited alternative sources of food and employment and small weak economies. Fisheries in inland water such as rivers, lakes, reservoirs and wetlands will be also strongly affected by changes in rainfall and increased temperatures. Even though the future of inland fisheries varies between continents, in Africa and Asia the resources are already very intensely exploited and are most at risk [24]. Vulnerable Asian countries have high fisheries dependence, heavily exploited marine ecosystems and a high exposure of coastal fisheries to changes in climate [7]. Much of Southeast Asia is seriously threatened by rising sea levels and changes in freshwater availability due to modifying snowmelt and precipitation. However, there are indications that production of pelagic fish in the North West Indian Ocean might increase as a result of climate change. Reduced snow cover across Eurasia is thought to be responsible for an increase in the strength of the southwesterly monsoon winds over the Arabian Sea. Wind strength stirs up deep water and increases the upwelling of nutrients into surface where it supports increased phytoplankton production. This availability of more nutrients at the base of the food web should boost pelagic production. In Latin American and Caribbean, high temperatures and increases in extreme weather events are expected to result in widespread loss of marine species and reduced availability and quality of freshwater as well as damage to coral reefs and mangroves [15]. In general, small-scale fisheries will be the first sectors to feel climate change impacts with consequences that include falling productivity, species migration and localized extinctions, as well as conflict over use of scarce resources. Therefore, the reduction in productivity will be negative for fishers at low latitudes where the majority of the world’s small-scale fisheries are located [3]. This could have significant effects on food security in those areas that are particularly vulnerable. In contrast, high latitude areas may experience localized increases in fish stocks due to immigration of species rising primary production. Governance will need to detect and control fishing capacity developments fast enough to reduce stress on locally declining species and let the new species settle successfully in new areas, avoiding overexploitation. Fish availability to consumers will depend on governance performance and eventual redirection of global trade flows and cannot be generically predicted. In any case, more research is needed into the impacts of climate change at the levels of local fisheries [14].

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12.5 Aquaculture food production Aquaculture contributed 43 % of aquatic animal food for human consumption in 2007 (e.g. fish, crustaceans and mollusks) and is expected to grow further to meet future demand [26]. The impacts of climate change on aquaculture are more complex than those on terrestrial agriculture owing to the much wider variety of species produced; however, it is different than fisheries because of the greater level of control possible over the production environment [16]. Aquaculture can provide food, control breeding, change environmental conditions by modifying water flows, temperature and water quality thus reducing dependence on ecosystem services. On the other hand, the construction of water storage and irrigation infrastructure may interfere with wild fisheries, but introducing aquaculture stocking can provide livelihoods for local communities and boost food security. However, climatic changes could affect productivity of aquaculture systems, increase the vulnerability of cultured fish to diseases and reduce returns to farmers (Tab. 12.1) Extreme weather events may result in the escape of farmed stock and contribute to reduction in genetic diversity of wild fish; this affects biodiversity. The mainstay of farmed fish production is freshwater omnivorous and herbivorous. This lies in marked contrast to capture fisheries in which the bulk of the fish species are marine carnivorous. However, many forms of aquaculture, particularly in developing countries, still depend heavily on wild stocks for food and seed [15]. Fish farmers may benefit from expansion of the areas where aquaculture is viable due to increased temperatures and rising sea levels. This is the case in currently cooler areas, such as those in more northerly latitudes; rising temperatures may result in increased growth rates and food conversion efficiencies, longer growing seasons, reduced cold water mortality and more suitable areas for aquaculture [16]. At the policy level, important questions exist about the priority given to conserving the environment versus the exploitation of natural resources for food production [26]. While richer nations in Europe may be able to offset reduced food production by increasing imports, the environmental impact is transferred to other countries where options or control are more limited.

12.6 Harmful algal blooms Climate change is expected to affect food safety, including the occurrence of natural toxins in seafood production; however, to date, quantitative estimates are scarce [27]. Primary producers constitute the basis of marine food webs but sometimes are harmful algal blooms (HABs). During the past few decades, there has been an apparent increase in the occurrence, frequency, magnitude, and duration of HABs worldwide [22]. The major functional groups of toxic phytoplankton are diatoms, dinoflagellates and cyanobacteria [28]. In the ocean, important harmful algae are diatoms from the genus Pseudonitzschia and species of dinoflagellates from the genera Alexandrium, Pyro-

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dinium, Gymnodinium, Karenia, Dinophysis, Prorocentrum or Gambierdiscus among others. In freshwater, the most important HABs are caused by certain species of cyanobacteria (blue green algae) from the genera Anabaena, Microcystis, and Aphanizomenon [29]. In both marine and freshwater systems, microalgae cells can occupy the water creating an anoxic barrier between the surface and deeper waters that may lead to benthic mortalities. Phycotoxins may cause direct heavy damage to fish, birds or mammals (behavior alterations, development impairments in early stages of life [30], abortion and premature birth of marine mammals [30] and even death). In some cases, filter-feeding shellfish ingests phytoplankton and the toxins may accumulate within their tissues with no obvious adverse effects [32]. These act as vectors for human intoxication either directly through the consumption of contaminated fish or shellfish, or indirectly through the toxin transfer via the food web to crustaceans and carnivorous fishes. Because phycotoxins are often tasteless, odorless, and heat and acid stable, if the fish or shellfish is contaminated, the normal food preparation procedures will not prevent intoxication [33–35]. For humans, exposure to HAB toxins also results from drinking contaminated water, inhaling toxic aerosol or by contacting contaminated water. Phycotoxins may cause respiratory and digestive problems, memory loss, seizures, lesions and skin irritation, or even fatalities in humans [32, 36, 37]. An effective shellfish monitoring system, which shuts down aquaculture sites when toxins exceed regulatory limits, has clearly prevented serious impact to human health. However, the closure of these sites has an adverse economic impact [38]. The relationship between toxic aquatic microalgae species and climate change has become a high profile and discussed topic in recent years. Physical and chemical conditions will affect phytoplankton in different ways. The higher tolerance of natural populations to environmental factors might be due to the strain variability between or within populations [39]. Research is now focusing on the possible future impacts of changing hydrological conditions on harmful algal bloom (HAB) species and their toxicity around the world.

12.6.1 Impact of temperature change on harmful algal blooms Global average temperature is expected to rise as a result of climate change, thus creating a marine environment particularly suited to phytoplankton. Due to its influence on molecular kinetic energy, temperature acts directly on cell physiological processes and determines metabolic rates. Moderate increases in temperature should enhance photosynthesis and phytoplankton growth; particularly harmful warm water species thrive at elevated temperatures [39]. Therefore areas undergoing rapid warming may be among the most vulnerable to increased HABs. Since 1990, areas of the North Sea and North-East Atlantic have been warming and warm water phytoplankton species abundance have been increasing. In relation to that, the distribution of several HAB-forming dinoflagellates in the North-East At-

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lantic has been studied. It was found that bio-geographical boundary shifts in phytoplankton populations made possible by climate change have the potential to lead to the poleward spread of HAB species normally suited to milder waters [40]. This impact propagates up the food web through copepod herbivores to zooplankton carnivores because of tight trophic coupling. Future warming is therefore likely to alter the spatial distribution of primary and secondary pelagic production, affecting ecosystem services and placing additional stress on already depleted fish [41]. In addition to a general spread in phytoplankton distributions due to oceanic warming, this will also influence in HABs frequency in some regions. In the eastern North Sea, for example, the occurrence of exceptional HABs has tripled since 1980. This increase is almost certainly related to the warming up of the sea and decreased salinity, the variability in wind speed and changes in inflow from the North Atlantic to the North Sea that would result in a slightly earlier spring bloom. The distribution of different phytoplankton groups suggests an increase in the abundance of dinoflagellates, whereas diatoms, flagellates and Phaeocystis sp. remain comparable to current levels, or are decreasing. The swimming ability of dinoflagellates allows them to move below the upper stratified layer of the water column to find nutrients in the deeper layer. Therefore it is expected to be favored over other phytoplankton in marine environments under climate scenarios where the surface becomes depleted of nutrients required for growth [29]. It is hypothesized that blooms of dinoflagellates such as Dinophysis spp. may occur more frequently in the North Sea by 2040. Warmer temperatures may result in expanded warm water HAB species. One example is the tropical marine dinoflagellate, Gambierdiscus toxicus, associated with ciguatera fish poisoning. Additionally, several dinoflagellates may quickly respond to changes in the environment because they form resting cysts during their lifecycles [42]. Some cysts can remain viable for tens of years, but if growth conditions occur more frequently, cysts may be expected to reseed more often; this causes HABs [43]. There are also interesting climate-ocean interactions, which affect HABs. For instance, El Niño/Southern Oscillation and the Pacific Decadal Oscillation (ENSO and PDO, respectively) have warm and cool phases that typically last for six to 18 months and 20 to 30 years, respectively. During warm phases, sea surface temperatures in the eastern and equatorial Pacific Ocean are anomalously warm, stratification is enhanced, and upwelling of nutrient-rich water along the eastern Pacific coast is reduced [29]. This may lead to an increase in growth rates of some HAB taxa including Gymnodinium, Prorocentrum and Dinophysis spp. [44]. The important role that the wind-driven upwelling at coastal margins plays in the development of some HABs was already studied in northwest Spain. Both the intensity and duration of upwelling has decreased since 1966. This significantly increased the water renewal time of the Rias Baixas, which increased the number of days annually that mussel rafts are closed to extraction due to HABs consisting primarily of the marine dinoflagellate Dinophysis, which produce the toxin okadaic acid that causes the syndrome diarrheic shellfish poisoning (DSP) in humans [29].

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12.6.2 Acidification of waters and effect on harmful algal blooms There is no complete evidence linking HABs to ocean acidification, but it is possible that it causes changes in the phytoplankton community composition. The current scenario in global climate change and increased CO2 concentrations can greatly influence the microbial communities in various environments, particularly in aquatic ecosystems [45]. Increased concentrations of free CO2 could potentially favor photosynthesis and growth, since CO2 is the primary substrate for photosynthesis [39]. Changes in dissolved seawater CO2 and HCO−3 concentrations may result in enhanced growth of certain species but in inhibited growth of calcifying phytoplankton due to the dissolution of their biogenic calcium carbonate (CaCO3 ) shells [29]. A more acidic environment would favor, among others, the dinoflagellates. It is widely expected that increasing acidification will reduce calcification of other marine organisms such as corals. Therefore ocean acidification will slow the rebuilding of coral reefs and weaken their structure. However, recent research has found increased calcification in the phytoplankton species Emiliania huxleyi [46] while the calcification of Mediterranean corals appears to be little affected by carbon dioxide levels and pH. In freshwater systems, a pH decrease would reduce the incidence of CHABs. It was empirically demonstrated that freshwater cyanobacteria species associated with HABs are poor competitors with other phytoplankton at low pH. However, it is unknown if this will counter the predicted increase in growth rates of cyanobacteria in response to warmer temperatures and nutrient over enrichment of waters [29].

12.6.3 Impact of sea-level rise and increased precipitation on harmful algal communities In a warmer climate, precipitation will increase due to a more intense hydrological cycle [47]. Increased precipitation and flash flooding release sudden nutrient-rich water to coastal seas and changes salinity resulting in alterations to phytoplankton community composition and the occurrence of HABs. The nitrogen (N) and phosphorus (P) concentrations increased because of persistent river flow due to heavy precipitation. These changes induced a dramatic increase in microalgal biomass with a decreasing diatom-dinoflagellate ratio, and exacerbated harmful algal blooms [48]. Several HABforming species appear to be responding particularly well in regions that are both warming and becoming increasingly fresh. In this case, large dinoflagellates and filamentous cyanobacteria dominate versus chain-forming diatoms [48]. For example, studies on the Norwegian coastal waters of the North Sea indicate a decrease in salinity related to increased precipitation and increased terrestrial run-off. In this region, there is an abundance of HAB-forming species such as Dinophysis spp., Protoperidinium and Prorocentrum spp. [49].

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The inexorable increase in the global supply of major nutrients could drive changes in the productivity and composition of the phytoplankton communities not only in marine but also in freshwater ecosystems [50]. For example, domoic acid production by the globally distributed diatom genus Pseudonitzschia is regulated by nutrient availability. Also, some experiments demonstrate a strong synergism between high CO2 and silicate limitation and the increase Pseudonitzschia cellular toxicity [51].

12.6.4 Microalgal toxicity As of today, a clear relation between climate changes and dinoflagellate toxicity is still speculative. Nevertheless, toxin production may be related to hydroclimatic conditions. For instance, Alexandrium ostenfeldii blooms may expand and increase in toxicity under high water temperature and atmospheric CO2 conditions [39]. Among other factors, nutrient limitation could enhance specific toxicity of algae due to an imbalance in toxin production and dilution through growth. Indeed, for Dinophysis acuminata, the highest cellular toxin content has been related to nitrogen limitation. Prorocentrum lima, a dinoflagellate with similar toxin profile, accumulated most toxin at low phosphate availability [49]. In general, phosphorus limitation causes an increase in the N-rich toxin (saxitoxin, cylindrospermopsin, microcystin and nodularin). Limitation by either nitrogen or phosphorus promotes the C-rich toxin (anatoxin, domoic acid, gymnodimine, spirolide, palytoxin, okadaic acid, brevetoxin, karlotoxin, ciguatoxin and maitotoxin) [52]. These observed relationships may assist in predicting and managing toxin-producing phytoplankton blooms. Generally, the response of Nand C-rich toxins to light availability seems less pronounced as compared to nutrient limitation, while enhanced CO2 availabilities may further promote the production of C-rich toxins under nutrient limiting conditions. The relationship between phytoplankton biomass and HAB species toxicity is complicated. Some HABs can have harmful effects even if the toxic species are not dominant, while others species are only toxic at high concentrations [52]. Also toxicity can vary among strains within a species, even during the course of one bloom event. Furthermore, it was suggested that a relationship exists between some bacteria and algal toxin production; therefore, if bacteria growth increases with climate change, the toxicity of some HAB species might also increase [53]. However, implications for shellfish toxicity remain unclear [49].

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12.7 Harmful algal blooms and aquatic food safety It is important to assess the potential effects of toxic HABs in aquatic food safety. Changes in climate may be creating a marine environment particularly suited to the increase in the occurrence of phytoplankton blooms. Therefore HABs were estimated to occur more often, but consequences for contamination of shellfish with toxins are uncertain. Changes in the occurrence of natural toxins are expected to affect the safety of food products arising from aquatic production systems [54]. These toxins are chemically stable substances that can have harmful effects to animals and humans after consumption of contaminated food products. Quantitatively estimating toxicological effects is difficult due, for instance, to the dearth of literature concerning the epidemiology of these toxins on human health [22, 55]. The ingestion of seafood contaminated with natural toxins produced by HAB organisms caused a number of human illnesses. These include amnesic shellfish poisoning (ASP), azaspiracid shellfish poisoning (AZP), ciguatera fish poisoning (CFP), cyclic imines poisoning (CIP), diarrheic shellfish poisoning (DSP), neurotoxic shellfish poisoning (NSP), palytoxin poisoning (PLP), paralytic shellfish poisoning (PSP), pectenotoxin poisoning (PeP), yessotoxin shellfish poisoning (YPS) [33, 36]. Some of the toxins responsible for the poisonings can be acutely lethal; in addition, no antidote exists for them [32, 56]. In addition to human health effects, HABs also have detrimental economic impacts due to the closure of commercial fisheries, public health costs and other related environmental impacts [57]. Amnesic shellfish poisoning (ASP). Pseudonitzschia species have the capacity to produce the potent neurotoxin domoic acid (DA), sometimes generating massive toxic HABs in coastal waters [58]. In addition to mussels, domoic acid (DA) can enter the food chain through vectors such as scallops, razor clams and crustaceans [59]. There have been many worldwide reports of DA contamination of seafood and mortalities to marine animals and birds. Live bivalve mollusks placed on the market for human consumption must not contain this marine biotoxin in total quantities (measured in the whole body or any part edible separately) that exceed the limit of 20 mg of DA per kilogram (Regulation (EC) No. 853/2004) [60]. Although there are many analytical methods for the determination of DA in seafood, liquid chromatography with ultraviolet detection is used by most regulatory agencies. Ciguatera fish poisoning (CFP). The marine benthic dinoflagellate of the genus Gambierdiscus produces ciguatoxins (CTXs) as secondary metabolites. CTXs cause ciguatera fish poisoning (CFP). This complex syndrome produces a wide variety of gastrointestinal, neurological and cardiovascular symptoms beginning a few hours after ingestion of contaminated seafood [61]. The identification of tropical and subtropical representatives of potentially toxic benthic dinoflagellate genera, such as Ostreopsis and Gambierdiscus, in the Eastern Atlantic Ocean has posed the question of the con-

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tribution of climate change on the spreading of microalgae to more temperate waters and its impact on food safety and human health [62]. About the possible risk associated to ciguatera, toxicity studies in the western Atlantic showed that in the Northern hemisphere, Gambierdiscus strains located at higher latitudes (Bahamas) were less toxic than those at lower latitudes (Caribbean). CTXs are transmitted through the food web starting with dinoflagellates before moving on to herbivorous fish, and then to the carnivorous fish [63]. Some of these toxins accumulate in the fish tissue and can be metabolized into the different forms (e.g. CTX-1B and 51-hydroxy CTX-3C) from original CTX structures (e.g. CTX-3C and CTX-4B) that are ultimately responsible for human intoxication [62, 64]. Therefore, CFP is a foodborne disease contracted by the consumption of finfish that have accumulated the lipid-soluble toxins CTXs in their flesh and viscera [65]. Carnivorous fish, and those situated at higher trophic levels within the food webs, can be considered riskier than those situated at lower levels (herbivorous). However, the complex transfer of toxins through the food webs and fish species involved does not always allow a clear determination of which fish species should be avoided. Additionally, large individuals are generally more toxic than small ones, since ciguatoxins accumulate in fish via the food chain [66]. On the other hand, diversity of fish species as toxin vectors may also be altered in response to different climatic conditions [67]. Some of these fishes, which become toxic in ciguateric areas, could expand CFP to higher latitudes by their migration due to global warming, although Gambierdiscus spp. is absent from these areas. Extrapolation of this model suggest the importance of monitoring sea surface temperature in addition to Gambierdiscus spp. populations for the assessment of a possible onset of ciguatera in the Mediterranean Sea and Eastern Atlantic Ocean in relation to global warming [62]. For a number of marine toxins, data from human intoxication incidents and also toxicological data obtained in laboratory tests have contributed to help regulatory authorities set a Maximum Permitted Level (MPL) for each of these toxins. Above this MPL, seafood must be banned from the market. For instance, MPL 0.01 ng g−1 P-CTX-1 equivalent toxicity was proposed for fishery products caught in the Pacific in 2000, and an MPL of 0.1 ng g−1 C-CTX-1 has been proposed for Caribbean fish [68, 69]. Currently there are no regulatory limits for CTX-group toxins in the European Union (EU), but the European Food Safety Authority (EFSA) states that checks are to take place to ensure that fishery products containing ciguatoxin are not to be placed on the market [61]. Interestingly, and contrary to other microalgal toxins implicated in food poisoning, no recognized official method for the determination of CFP toxins exists [62]. However, the mouse bioassay has been widely used in the past for screening ciguatoxicity in fish samples [70]. It is necessary to have adequate CTX quantification methods to diagnose CFP cases and to prevent intoxications through the analysis of consumable fish. Therefore, improving detection of CTXs mainly through the establishment of monitoring programs for toxins in food and understanding toxin dinoflagellate and toxin vector spatial distribution will contribute to improving the risk analysis of CFP in the world. With this strategy in mind, there should be links between food safety and medical care.

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Diarrheic shellfish poisoning (DSP) is a specific type of food poisoning, characterized by severe gastrointestinal illness due to the ingestion of filter-feeding bivalves contaminated with okadaic acid (OA) or dinophysistoxins (DTXs) [71, 72]. Highest risk areas are those combining Dinophysis strains with high cell content of okadaates, aquaculture with predominance of mytilids as well as consumers who frequently include mussels in their diet [73]. The increase in HAB frequency and expansion related to DSP have led most countries to establish regular activities for both monitoring harmful phytoplankton in coastal waters and for assessing levels of toxins in farmed mollusks. The acute reference dose (ARfD) established by the European Food Safety Authority (EFSA) is 0.3 mg OA equivalents/kg bw [74]. Therefore, live bivalve mollusks placed on the market for human consumption must not contain those marine biotoxins in total quantities that exceed the limit of 160 μg OA equivalents per kg [75]. This is the case even though, based on acute effects on consumers, a decrease of this level to 45 μg/kg was already proposed in a recent report by the European Food Safety Authority (EFSA) [74]. Those toxins are produced by Dinopysis spp [33]. Climate change could influence species composition within Dinophysis spp. which would affect the toxin production per cell. If the group of Dinophysis spp. behaves similarly to other dinoflagellates in the future, then the frequency of harmful algal blooms of Dinophysis spp. may also increase, but consequences for contamination of shellfish with diarrheic shellfish toxins are uncertain [27]. Neurotoxic shellfish poisoning (NSP). Brevetoxins are responsible for NSP and can cause significant mortalities of fish and other aquatic animals, including marine mammals, through direct exposure during the marine dinoflagellate Karenia brevis blooms [76]. However, many filter-feeding mollusks are known to accumulate brevetoxins according to their natural food chain without adverse effects. Those mollusks can be a food source for other aquatic animals; therefore, bio-accumulation in the food web is occurring [77]. Like many marine toxins, depuration time of brevetoxins in shellfish varies, but is typically within two to eight weeks, although reports of much longer retention have been documented [78]. Brevetoxins are not diminished by rinsing, cleaning, cooking or freezing. The inability to easily detect the presence of brevetoxins (for instance by taste or smell) in these human food sources past bloom cessation potentially puts consumers at an increased risk for NSP. The ingestion of even a few contaminated shellfish may result in poisoning. The severity of the disease may be dependent upon many factors including dose, body weight, underlying medical conditions and age of the victim as well as possibly the toxin mixture of the particular bloom. Currently there are no regulatory limits for brevetoxins in shellfish or fish in Europe because to date they have not been reported [79]. However, environmental change will affect the distribution, duration and frequency of blooms of K. brevis and other brevetoxin producing algae in all aquatic environments [77]. Brevetoxins could also emerge in Europe. Besides, globalization of the food supply will require more effective and vigilant monitoring of algal-derived toxins in fish and shellfish.

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Palytoxin poisoning. Dinoflagellates of the genus Ostreopsis are known to cause food poisoning in tropical coastal areas following the accumulation of palytoxin (PLTX) and/or its analogues in crabs, sea urchins or fish. Over the last ten years tropical Ostreopsis spp.have become more frequent in the Mediterranean Sea (Italy, Spain, Greece, Tunisa, Monaco and France) and in the Eastern Atlantic Ocean probably in response to climate change [62, 80]. In this case, the toxins also appeared in mussels and the peaks of toxicity were consistent with the peak of cell concentrations of Ostreopsis spp. in the field [81]. During the blooms of Ostreopsis spp. recorded in 2008 and 2009 in the Mediterranean Sea, concentrations of total PLTX exceeded the threshold value for the protection of public health established in late 2009 by the European Food Safety Authority (EFSA): 30 μg PLTX/kg shellfish flesh [82]. Concerning food safety, such a threshold of health protection requires monitoring of the seafood contamination (sea urchins, mussels and fish), from commercial fishing and recreational fishing at bathing sites contaminated by Ostreopsis [80]. The consumption of seafood in the affected areas produced signs of direct poisoning in several people. Symptoms after PTX ingestion may include: myalgias, weakness and neuromuscular dysfunction, wheezing or respiratory distress possibly due to muscular contraction, delayed hemolysis and cardiac conduction abnormalities [83]. Indeed, the toxins can be found in sea spray droplets dispersed by the wind. Inhalation of contaminated spray was responsible for outbreaks of dramatic febrile respiratory syndromes and irritation of the skin and upper respiratory area [84]. PLTX is also found in soft corals from Palythoa that could be used in home aquariums and may produce dermal and respiratory intoxications. Paralytic shellfish poisoning. It is the most widespread of the HAB-related shellfish poisoning syndromes. Saxitoxin (STX) group toxins (produced by Alexandrium spp., Gymnodinium catenatum or Pyrodinum bahamense) are those responsible for paralytic shellfish poisoning. PSP results from ingestion of shellfish that accumulate toxins from feeding on Alexandrium cells vary. Symptoms range from a slight tingling sensation or numbness around the lips to fatal respiratory paralysis. For the control of the STX group toxins in seafood, the current EU regulatory level is 800 μg STX equivalents/kg shellfish meat [85]. The genera Alexandrium is one of the most studied marine dinoflagellate groups due to its ecological, toxicological and economic importance. Phenotypic variability within populations plays an important role in the adaptation of phytoplankton to changing environments, potentially attenuating short-term effects and forming the basis for selection. In particular, A. ostenfeldii blooms may expand and increase in toxicity under increased water temperature and atmospheric CO2 availability with potentially severe consequences for the coastal ecosystem and food safety [39]. Toxic Alexandrium species are found worldwide in temperate coastal and estuarine waters with recurrent blooms affecting economy, human health and ecosystem.

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Yessotoxin. Over the past 15 years, the Mediterranean Sea has been characterized by the presence of potentially toxic algae, with consequences for human or faunal health not sufficiently clarified. Some problems have been dealing with the occurrence of yessotoxin-producing species (i.e. Protoceratium reticulatum, Gonyaulax spinifera and Lingulodinium polyedrum) that prompted mussel farms closure for several months of many years [86]. Toxic and non-toxic strains of Lingulodinium polyedrum were also reported from different areas; it also seems that it only becomes toxic under certain environmental conditions [87]. So far, live bivalve mollusks placed on the market for human consumption must not contain those marine biotoxins in total quantities (measured in the whole body or any part edible separately) that exceed the limit of one milligram of YTX equivalent per kg shellfish meat [88]. However in 2013 EFSA CONTAM Panel (Panel on Contaminants in the Food Chain) suggested that for yessotoxins, the current limit of 1 mg/kg be increased to 3.75 mg/kg of shellfish flesh. Tetrodotoxin (TTX). Bacteria associated with TTX-bearing animals and their food produce TTX [89]. The occurrence of TTX has been mainly reported in Asian countries and more specifically in Japan, with many cases of TTX-food poisoning related to consuming pufferfish [90]. However, climate change causes the global distribution of tetrodotoxin that is spreading with the United States [91] and the European Atlantic coastline [92]. Moreover, accumulation, metabolism and detoxification kinetics of tetrodotoxins may be altered in fish vectors in response to the variation of ambient temperature. In the EU, there are severe limitations in the legislation in relation to TTX contamination in seafood. Regulation (EC) no. 854/2004 stipulates that: “Fishery products derived from poisonous fish of the following families: Tetraodontidae, Molidae, Diodontidae and Canthigasteridae must not be placed on the market” [60]. In Japan, the government has set regulatory limits for TTX in food of 2000 μg/kg TTX equivalents [34], whereas the United States has a zero tolerance level due to the fact that no product sold legally in the United States is expected to contain this toxin [93]. Microcystins (MCs). The cyanobacterial community may also be altered by various physical/chemical factors, including change in environmental conditions; it is essential to understand and predict the responses of these communities [45]. Microcystins (MCs) represent a group of 80 structural heptapeptide variants with hepatotoxicity, renal toxicity and neurotoxicity [94, 95]. MCs are produced by a large variety of planktonic and benthic cyanobacterial genera including Microcystis, Nostoc, Planktothrix, Anabaena, Synechococcus and Snowella [96]. Some authors postulated that MC synthesis is increased under optimal temperature and growth conditions [97]. Thus a warmer climate could elevate the temperature above the minimum threshold level for toxin synthesis, leading to an increased general metabolic activity and thus a high level of toxin production [98].

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Food safety. Over the past few decades, various types of poisoning incidents have been reported [99, 100] and this has led to growing concerns with regard to the consumption of seafood. Very few epidemiologic studies have been done that are designed to systematically assess the adverse health effects from exposure to HAB toxins. This will make the detection and quantification of climate change impacts on HABrelated illnesses difficult [57]. Probably the incidence of human syndromes associated with exposure to toxic seafood will increase as HABs occur more frequently and over greater geographic areas. It will be important to monitor any increases in the susceptibility of seafood species to contamination by HABs, as well as the subsequent risks to consumers. Public health, fisheries and aquaculture industries have been affected by poisoning associated to aquatic toxins [92]. Improved capacity to predict HABs is important for risk management, but other tools are needed as well. Governments have to ensure that integrated shellfish and microalgal monitoring programs are adequate and in line with international recommendations [2, 3, 15]. While these are not specific to climate change, the increased of biotoxin contamination events makes it worthwhile to focus attention on these requirements. It is recommended to closely monitor levels of marine biotoxins in the future, in particular related to risky situations associated with favorable climatic conditions for toxin-producing organisms [27]. Countries are encouraged to implement marine biotoxin management plans to strengthen risk management capability and to enhance consumer protection. Those countries should generate more toxicological data and improve and validate toxin detection methods in shellfish. However, developing countries often lack food safety standards employed in developed regions as well as the health care capacity to deal with outbreaks caused by shellfish or fish toxicity. FAO, in collaboration with many international, intergovernmental and governmental bodies, has supported the development of a standardized training programmed to assist countries in understanding and carrying out food safety assessments [2, 3].

12.7.1 Predictive modeling To date, quantitative estimates of climate change’s effect on food production and feed safety, including the occurrence of natural toxins in seafood, are scarce [27, 101]. HAB prediction is important for more effective risk management, but it is complex as it includes conceptual descriptions of ecological relationships and statistically based empirical as well as numerical models. There is much ongoing research in this sense to improve our knowledge of the factors that influence population dynamics of harmful algae [102]. Microalgal monitoring coupled with operational oceanographic, meteorological and remote sensing data, including modeling and other measurements are being used in the prediction of HABs. In addition, statistical data analyses using existing national datasets from the study area were performed to obtain information

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on the relationships between dinoflagellate cell counts and contamination of shellfish with toxins. The international dimension is also important for understanding and addressing the global impact of climate change on HABs. Now several national and international programs exist that study and monitor the ecology of HABs in marine and freshwater. Based on predictive modeling, it an expansion in area and frequency of HABs was projected along Northern Europe and Northern Asia. The implications are shifts in vulnerability of coastal systems to HAB events, increased regional HAB impacts to aquaculture, increased risks to human health and ecosystems, and economic consequences of these events due to losses to fisheries and ecosystem services [37].

12.8 Future perspectives Climate change is projected to have substantial effects on the frequency and abundance of HABs because of the complexity of factors that may affect the growth or habitat of microalgae. Those factors include temperature, altered salinity due to increased precipitation and runoff, increased stratification and changes in nutrient and light regimes [37]. Harmful algal blooms (HABs) can cause fish kills, contaminate seafood with toxins or detrimentally alter ecosystem functions [22]. If climate change were slow, adaptation would be easier [14]. Some aspects of climate change will be most pronounced at high latitudes; however, the countries with economies most vulnerable to those effects on fisheries are in Africa, northwestern South America and Asia. Understanding how climate exposure, fisheries dependence and adaptive capacity combine in those regions to influence vulnerability provides a useful starting point for directing future research and mitigation initiatives [7]. Fishers and fish farmers must adapt to climate change in ways that allow them to moderate potential damages or to take advantage of opportunities. Options include actions that increase the adaptive capacity of ecosystems by reducing environmental stresses (over-fishing, habitat destruction and pollution) that can significantly increase the vulnerability of communities [103]. The adaptation in aquaculture needed to develop new strains of species that are tolerant of lower water quality and different levels of salinity to cope with environmental changes. Assuring food safety is a complex task and ultimately depends on the ability of countries to cope with risks. Therefore governments must implement controls at the most appropriate point within the food production systems to present safe food to consumers. Several developed countries have already initiated programs of work aimed at identifying emerging food safety risks linked to climate change. The Food and Agriculture Organization (FAO) of the United Nations has a key role in assisting them to assess the changes to their food safety and to promote international cooperation in improving the understanding of food safety implications of climate change [3, 11, 15].

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Research is needed to evaluate the direct and indirect associations between climate change, HABs, food production and safety, and human health. It requires a combined oceanographic-epidemiologic study that addresses very long-term and extensive geographic scales. Monitoring and surveillance programs are necessary to handle emerging hazards arising from global climate change. The data generated from these progams will contribute significantly to predictive modeling and risk assessments and should be shared readily both at national and international level. Risk assessment will provide the scientific basis for the development and adoption of food safety measures. Finally, climate change may improve the conditions for some resources and worsen them for others. Unless climate change factors lead to major losses of aquatic productivity, e.g. through food chain disruptions, the global consequences of climate change on the world contribution of fish to food security might be minimal. However, local consequences could be rather serious, particularly on poor rural coastal areas, and would need to be further assessed [14]. Therefore, the fishery sector requires special consideration to ensure that policy responses to climate change are effective. An Integrated Maritime Policy will be a key tool to fight climate change and adapt to its impacts. The best situation would be if climate change was slow; adaptation would thus be easier and the impact on food production and safety would be minimal.

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Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana

13 From science to policy: dynamic adaptation of legal regulations on aquatic biotoxins 13.1 Introduction Historically, legal regulations have evolved with the appearance of new toxins or new locations of known toxins to protect human health. Although no direct link can be undoubtedly established between climate change and the dissemination of toxic episodes to new geographic areas, the fact is that new toxins or new locations seem to be emerging with a higher frequency in recent years. Many factors can influence the increment of reported episodes, such as monitoring intensification and technological development, but the contribution of global warming has also been widely discussed in the scientific community. Regulations and monitoring programs will have to be adapted to new toxicological threats worldwide, both for marine phycotoxins and cyanotoxins. In this chapter we review the current regulatory status of marine phycotoxins and cyanotoxins worldwide and discuss the problems that regulatory agencies will have in establishing and implementing new maximum limits.

13.2 Current worldwide regulations on marine phycotoxins 13.2.1 Maximum permitted levels Current worldwide regulations on the maximum content of marine phycotoxins in seafood are based mainly on the toxicological evaluation of these toxins by ad hoc experts designated by FAO, WHO and IOC, and published in a joint research report on marine biotoxins in 2004 [1, 2]. EFSA panels of experts on several toxin classes have also contributed important information for the regulation of these toxins worldwide [3–7]. The aim of the regulations on marine biotoxins is the protection of the human population, and therefore an important contribution of these texts is an international group of experts’ evaluation of the toxicological data available for several toxin classes, and the subsequent suggestion of a LOAEL (lowest observable adverse effect level) for every class and a derived acute reference dose (ARfD). An important limitation for many toxins is the scarcity of toxicological data both in humans or animals to make an adequate evaluation of the toxic risk, and therefore the values of ARfD proposed by these working groups are usually considered provisional until more information is available. Unfortunately, a tolerable daily intake (TDI) for these toxins,

442 | Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana

which would be a better parameter to warrant human safety, cannot be calculated due to the lack of chronic toxicity studies. LOAELs and the derived ARfDs proposed in these reports are shown in Tab. 13.1. Tab. 13.1: LOAEL and ARfD values proposed by the joint FAO/WHO/IOC report and the EFSA panels. LOAEL NOAEL μg/kg of BW

Safety factor

ARfD μg/kg of BW

Toxin group

Report

AZA

FAO/WHO/IOC

0.4

10

0.04

EFSA

1.9

3

0.2

Derived guidance level based on consumption* (portion size) 24 μg/kg 9.6 μg/kg 6.3 μg/kg 30 μg/kg

of SM (100 g) of SM (250 g) of SM (380 g) of SM (400 g)

BTX

FAO/WHO/IOC EFSA

— —

— —

— —

— —

CTX

FAO/WHO/IOC EFSA

— —

— —

— —

— —

Cyclic imines

FAO/WHO/IOC EFSA

— —

— —

— —

— —

DA

FAO/WHO/IOC

1000

10

100

900

30

30

60 mg/kg 24 mg/kg 16 mg/kg 4.5 mg/kg

of SM (100 g) of SM (250 g) of SM (380 g) of SM (400 g)

FAO/WHO/IOC

1

3

0.33

EFSA

0.8

3

0.3

200 μg/kg 80 μg/kg 50 μg/kg 45 μg/kg

of SM (100 g) of SM (250 g) of SM (380 g) of SM (400 g)

0.2

30 μg/kg

of SM (400 g)

0.8

120 μg/kg

of SM (400 g)

420 μg/kg 170 μg/kg 110 μg/kg 75 μg/kg

of SM (100 g) of SM (250 g) of SM (380 g) of SM (400 g)

EFSA OA

Palytoxin PTX STX

YTX

FAO/WHO/IOC EFSA

— 200

— 1000

—

FAO/WHO/IOC EFSA

— 250

— 300

—

FAO/WHO/IOC

2

3

0.7

EFSA

1.5

3

0.5

FAO/WHO/IOC

5000

100

50

EFSA

5000

200

25

30 mg/kg of SM (100 g) 12 mg/kg of SM (250 g) 8 mg/kg of SM (380 g) 3.75 mg/kg of SM (400 g)

* Derived guidance level to avoid exceeding ARfD calculated based on an adult BW of 60 kg. BW: body weight, AZA: azaspiracid, BTX: brevetoxin, DA: domoic acid, PTX: pectenotoxin, SM: shellfish meat, STX: saxitoxin, YTX: yessotoxin. References: [1–6, 8–13].

OA: okadaic acid,

13 From science to policy: dynamic adaptation of legal regulations | 443

An assessment of the risk of exposure to these toxins is also required for the calculation of the maximum concentration that should be allowed in seafood. For this purpose, dietary intake of each food class should be considered [1, 2, 8, 9]. Most differences among countries regarding regulatory levels of toxins have been related to variations in the estimation of the portion size. Actually, diet composition may vary greatly among countries. The FAO/WHO/IOC joint report provides guidance levels based on three portion sizes of shellfish meat (100, 250 and 380 g; Tab. 13.1), although the recommendation is to use 250 g portions. EFSA recommendations are calculated on a 400 g portion size basis [1–6, 8–13]. As a result of this huge international effort, Codex Alimentarius has published provisions on the maximum levels of several marine biotoxins in bivalve mollusks in the “Standard for live and raw bivalves” [14]. This document, revised in 2014, establishes maximum toxin levels in shellfish for the saxitoxin (STX), okadaic acid (OA), domoic acid (DA), brevetoxin (BTX) and azaspiracid (AZA) groups (Tab. 13.2). Most countries’ current regulations follow the Codex maximum levels for these toxin classes (see Tab. 13.2 for examples). For the toxin groups not mentioned in the Codex standard, the regulations are not homogeneous (Tab. 13.2). Yessotoxins and pectenotoxins are regulated in some countries, while others do not consider that there is enough evidence to demonstrate that they are a hazard to consumers because no human illness has been linked to exposure to these toxins [11, 12, 15]. Ciguatoxins (CTXs) are also regulated in some countries. In Australia, where ciguatera is endemic in the north-eastern regions, reef fish species known to be highrisk ciguatera vectors are not allowed to enter the market depending on their size and capture location [16]. A list of susceptible species and size limits is available in the Sidney Fish Market guidelines [17]. In addition, the capture of barracudas and Spanish mackerel is banned by the Queensland Government in Platypus Bay, a well-known source of ciguateric fish, and clinical cases of ciguatera must also be reported to health authorities [16]. In the USA, action levels for Pacific and Caribbean CTXs have been established in consideration of their toxic potencies [15], while in Mexico there is no distinction between both classes of CTXs, with just a general value for the group [18] (Tab. 13.2). The FDA has also published guidelines for purchasing fish species that are vectors of ciguatera [19]. In Europe, fishery products containing CTXs are not allowed on the market [20]. In Japan, the import and sale of fish species known to be ciguatera vectors is banned or conditionally allowed; a limit has also been established (Tab. 13.2) [6, 21]. Puffer fish harvest and imports are usually banned or absence of tetrodotoxin (TTX) in fish is required [15, 20, 22]. In Japan, regulations permit certain fishing areas, species of puffer fish, edible parts and a maximum TTX level [21]. Preparation establishments and licensing of fugu chefs is also regulated.

444 | Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana

Tab. 13.2: Legislation on marine phycotoxins presence in seafood in seven countries and Codex provisions. Country

Toxin

Food product

Maximum limit

CODEX

STX

Mollusks

800 μg STX eq./kg

DA

Mollusks

20 mg/kg

Lipophilic toxins OA, DTXs AZAs

Mollusks

BTX

Mollusks

200 MU/kg

PSP

Bivalve mollusks

800 μg STX eq./kg

Mouse bioassay or HPLC-FLD

DA

Bivalve mollusks

20 mg/kg

LC-MS or HPLC

Lipophilic toxins (OA and DTXs)

Bivalve mollusks

200 μg OA eq./kg

LC-MS/MS or HPLC-MS

NSP

Bivalve mollusks

200 MU/kg (≈ 800 μg BTX-2 eq./kg)

MBA or LC-MS

CTX

Fish

Ban on sale of high risk species dependent on size limits and capture location

PSP toxins

Shellfish

800 μg STX eq./kg

MBA

DA

Shellfish

20 mg/kg

HPLC-UVD

Lipophilic toxins OA, DTXs, PTXs

Shellfish

200 μg OA eq./kg

MBA

STX

Mollusks

800 μg STX eq./kg

MBA

DA

Mollusks

20 mg/kg

HPLC-UVD

Lipophilic toxins OA, DTXs, PTXs YTXs AZAs

Mollusks

PSP toxins

Bivalve mollusks

800 μg STX eq./kg

MBA (R) HPLC-FLD (alternative)

DA

Bivalve mollusks

20 mg/kg

HPLC-UVD

Lipophilic toxins OA, DTXs, PTXs YTXs AZAs

Bivalve mollusks

Australia

Canada

Chile

European Union

Official method/ accepted (R): reference

160 μg OA eq./kg 160 μg OA eq./kg

MBA 160 μg OA eq./kg 3.75 mg/kg 160 μg OA eq./kg

LC-MS 160 μg OA eq./kg 3.75 mg/kg 160 μg OA eq./kg

13 From science to policy: dynamic adaptation of legal regulations | 445

Tab. 13.2 (continued) Country

Japan

Mexico

New Zealand

USA

Toxin

Food product

Maximum limit

Official method/ accepted (R): reference

TTX

Fish

Absence

—

CTX

Fish

Absence

—

PSP toxins

Bivalve mollusks

800 μg STX eq./kg

Lipophilic toxins DSP (OA group) AZAs

Bivalve mollusks

MBA MBA, LC-MS

160 μg OA eq./kg 160 μg OA eq./kg

DA

Bivalve mollusks

20 mg/kg

LC-MS

NSP

Shellfish

20 MU/100 g

MBA, LC-MS

TTX

Puffer fish

10 MU/g. Restricted capture and sale

MBA

CTX

Fish

0.025 MU/g. Banned import and sale of ciguateric fish

MBA

STX

Mollusks

800 μg STX eq./kg

MBA

DA

Mollusks

20 mg/kg

HPLC-UV

BTX

Mollusks

20 MU/100 g

MBA

Lipophilic toxins OA, DTXs

Mollusks

CTX

Fish from tropical and subtropical areas

MBA 160 μg OA eq./kg 2.5 MU/100 g

MBA

PSP toxins

Bivalve mollusks

800 μg STX eq./kg

LC-MS, MBA

DA

Bivalve mollusks

20 mg/kg

HPLC-UVD, LC-MS

Lipophilic toxins OA, DTXs, PTXs YTXs AZAs

Bivalve mollusks

BTX

Bivalve mollusks

20 MU/100 g (≈ 800 μg BTX-2 eq./kg)

LC-MS, MBA

PSP toxins

Mollusks Crustaceans Finfish

800 μg STX eq./kg

HPLC-FLD MBA RBA

DA

Mollusks Crustaceans Finfish

20 mg/kg

HPLC-UVD

LC-MS 160 μg OA eq./kg 1 mg/kg 160 μg OA eq./kg

446 | Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana Tab. 13.2 (continued) Country

Toxin

Food product

Lipophilic toxins DSP toxins AZAs

Bivalve mollusks

NSP

Bivalve mollusks

CTXs Pacific CTX Caribean CTX

Finfish

Tetrodotoxin

Puffer fish

Maximum limit

Official method/ accepted (R): reference HPLC-MS/MS

160 μg OA eq./kg 160 μg AZA-1 eq./kg 800 μg BTX-2 eq./kg

MBA —

10 ng/kg P-CTX-1 eq. 100 ng/kg P-CTX-1 eq. Ban on harvesting in Florida, restriction on importation of all species

HPLC-FLD: high performance liquid chromatography coupled to fluorescence detection, HPLC-UVD: high performance liquid chromatography coupled to ultraviolet detection, LC-MS: liquid chromatography coupled to mass spectrometry, MBA: mouse bioassay, MU: mouse units, PIA: phosphatase inhibition assay, RBA: receptor-binding assay. References: [15, 18, 20–22, 28, 34, 39, 117, 198–203].

13.2.2 Official detection methods In some countries, official and/or reference methods for the detection of regulated marine biotoxins are legislated. Official detection methods for PSP toxins are the mouse bioassay (AOAC 959.08), HPLC coupled to fluorescent detection (HPLC-FLD) with pre-column derivatization (Lawrence method, AOAC 2005.06) and with postcolumn oxidation (PCOX method, AOAC 2011.02) and the receptor-binding assay (AOAC 2011.27) [22–30]. For the detection of lipophilic toxins, the mouse bioassay and LC-MS/MS detection are currently being used in several countries following the standard operating procedures (SOPs) of the European Union Reference Laboratory on Marine Biotoxins (EU-RL-MB) [31, 32] or the European Committee of Standardization method EN16204:2012 [33]. Recently, the mouse bioassay has been replaced by LC-MS/MS as the official method for the detection of lipophilic toxins in the EU [34]. For a long time the official method for the detection of DA has been performed using HPLC coupled to UV detection (HPLC-UVD) [22, 35–37]. Currently, an immunoassay has also been published as an official method by the AOAC (AOAC 2006.02) [38] and included in several regulations [39], and the EU-RL-MB has also produced a SOP for UPLC-MS detection of DA [40]. Legislations of most countries do not mention official methods for the detection of regulated toxins; when they do, there is no uniformity among countries in the selection of official, reference, or accepted methods (Tab. 13.2). The “Standard for live and raw bivalve molluscs” (revised in 2014) includes a section

13 From science to policy: dynamic adaptation of legal regulations | 447

on methods for the determination of biotoxins [14]. This document establishes the minimum numerical criteria that the chemical methods for detection of marine toxins should meet in terms of LOD (limit of detection), LOQ (limit of quantification), precision and recovery. The techniques mentioned above are listed as appropriate methods for the detection of these marine toxins in the draft performance criteria for reference and confirmatory methods [41]. When toxicity is evaluated using analytical methods, the amounts of toxins detected should be transformed to toxicity values. This transformation should be based on the toxic potency of different analogs within a group. Therefore, a toxicity equivalency factor (TEF) is estimated for every analog considering its relative toxic potency against the representative toxin of the class. The final results of analytical methods are reported as the amount of representative toxin equivalents. There are disparities in the TEFs used by different laboratories, which obviously influences analytical results. Currently, internationally accepted TEF values are available in the FAO website for PSPs, DSPs and AZAs, although they are considered provisional [14, 41].

13.3 Current worldwide regulations on cyanotoxins Regulations on cyanotoxins in many countries followed WHO recommendations published in the Guidelines for Drinking-Water Quality in 1998 [42]. This document recommended a provisional 1 μg/l maximum content of microcystin-LR (MC-LR) in drinking water. MC-LR is probably the most common cyanotoxin, but not the only one that may appear in drinking waters. In addition, human exposure by the oral route may occur from recreational waters due to swallowing while swimming or from ingestion of contaminated freshwater fish or shellfish in addition to drinking water [43–46]. Therefore legal regulations of many countries have become more complex than this first recommendation as knowledge about cyanotoxins and their associated risks has expanded. The WHO provisional recommended level was based on an estimation of a tolerable daily intake (TDI) for humans of 0.04 μg MC-LR/kg/day. This TDI was calculated from the no observed adverse effect level of 40 μg/kg obtained in mice, applying several safety factors that accounted for inter- and intra-species variations and possible chronic effects [47]. The final guideline value was obtained considering an average body weight of 60 kg, a daily consumption of water of 2 l and a proportion of daily intake of 0.8. Further WHO recommendations for recreational waters were also based on this TDI value [48]. However, the toxicological data available for an adequate estimation of toxic risks were scarce at the time, and therefore the guideline clearly defines this recommended level as provisional. Most countries have recognized that the toxic threat from the microcystin group is not restricted to MC-LR. There are more than 80 compounds in this group and current regulations usually refer to microcystins and establish a regulatory limit in 1 μg of MCLR equivalents/l, but rarely specify the microcystin analogs that should be monitored.

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A few countries also consider the presence of cylindrospermopsin, STX and anatoxin-a providing guidance on maximum recommended levels. The regulations on cyanotoxin presence in drinking water in several countries are compiled in Tab. 13.3 (for further review, see [47, 49]). Control of human exposure to these toxins through drinking water has also prompted the development of monitoring programs of cyanobacteria growth in freshwater reservoirs. Water sampling protocols and alert systems have been designed to define intervention at different levels of cyanobacterial biomass or toxin presence, from intensification of monitoring to treatment or banning of use. Examples of these plans implemented in Australia, Brazil, the Czech Republic, Finland or New Zealand can be found in Chorus 2012 and Ibelings et al. 2014 [47, 49]. Frequency of sampling varies greatly among countries, from once a year to weekly, and in some cases water is only analyzed if there are signs of cyanobacterial blooming. Usually, signs of cyanobacterial growth trigger the switch from surveillance level to alert level and sampling frequency is incremented to at least once a week [49]. Regulations on recreational waters are usually based on cyanobacteria biomass and they are shown in Tab. 13.4. An additional factor that should be considered for protecting human health is the fact that cyanobacteria cells frequently form scums on the surface of bodies of water, which are often pushed to the shore. Mainly for intracellular toxins like microcystins, this phenomenon increases toxicity and exposure in areas where these cyanobacteria scums are accumulated. Recently benthic cyanobacterial growth of toxigenic species has also been recognized as a potential threat, and therefore monitoring of benthic cyanobacterial mats has also been regulated in two countries (Tab. 13.4) [47, 50]. The guidelines published by WHO for managing recreational water safety suggest monitoring cyanobacterial cells or chlorophyll-a levels and also visual inspection of scum presence (Tab. 13.4) [48]. The Bathing Water Directive of the European Union regulates the quality of waters used for recreational purposes in the EU; however, no specific parameters of water quality are legislated [51]. Therefore, specific regulations of different European countries present variations in the way this directive is being implemented, although WHO recommendations are followed most of the time (Tab. 13.4). On thing is worth noting: the regulations that control the presence of cyanotoxins in drinking water are fairly uniform worldwide, with a few small variations based usually on the average weight of population considered for the limit estimation or volume of water consumed [47]. However, the regulations on recreational waters differ greatly among countries with regards to criteria that trigger alert levels. A more restrictive protection of human health in some countries may be the origin of these differences, but probably the variations of the ecobiology of cyanobacterial growth and cyanotoxin production among countries is also a factor influencing these discrepancies. Finally the amount of cyanotoxins in food collected in freshwaters is only regulated in a few countries (Tab. 13.5).

Phytoplankton Cyanobacteria Phytoplankton and cyanobacteria fraction

Cuba (no regulations, framework being tested)

Czech Republic

Visual inspection monthly Sampling 4 times a year

< 20 × 103 cells/ml (SL) < 1.5 × 103 cells/ml (SL) 20–100 × 103 cells/ml and > 50 % cyanobacteria (AL 1) Toxic cyanobaceteria sp. (AL 1) Humans or animals (AL 2)

MC-LR ANAT-a

Canada

MC-LR Cyanobacteria

Report of toxic effects Scum presence

Dependent on bloom occurrence. ANAT-a only in Quebec

1.5 μg/l 3.7 μg/l

MCs CYL STX Cyanobacteria

1 μg/l ≥ 2 × 103 or 100 × 103 cells/ml (AL 1, 2 or 3 resp.) ≥ 0.2 or 10 mm3 /l biovol. (AL 1 or 2 resp.) ≥ 1 or 10 μg/l chlorophyll-a. (AL 1 or 2 resp.)

1 μg/l 15 μg/l 3 μg/l (STX eq.) 10–20 × 103 cells/ml or 1 mm3 /l biovol.

Once per week

Dependent on blooming

Dependent on risk of blooming for each area

Brazil

1.3 μg/l MC-LR eq. — 1 μg/l 3 μg/l (STX eq.)

MCs NOD CYL STX

Minimum Frequency

Australia

Maximum limit

Toxin legislatedRecommendation

Country

Drinking water Toxin concentration in finished drinking water unless otherwise stated Cyanobacteria concentration in water reservoir

Tab. 13.3: Legislation on cyanotoxins presence in drinking water.

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MCs

Finland

1 μg/l 0.1 μg/l —

—

1 μg/l < 4.7 × 103 cells/ml

MCs CYL

Cyanobacteria

Algae

MCs Cyanobacteria

MCs NOD CYL ANAT-a HomoANAT-a ANAT-a(s) STX

Germany (no specific regulations)

Hungary (no specific regulations)

Italy (no specific regulations)

Netherland (no specific regulations)

New Zealand

1 μg/l MC-LR eq. 1 μg/l 1 μg/l 6 μg/l 2 μg/l 1 μg/l 3 μg/l STX eq.

1 μg/l

Fortnight if area susceptible of cyanobacteria blooms

If presumed risk

At least once a year

Only when cyanobacteria proliferate

Visual inspection weekly

> 1 or 10 μg/l (restriction or ban on water use, AL 3 or 4 resp.) > 1 μg/l in raw water (AL 2) > 5 × 103 cells/ml or ≥ 1 mg/l biomass (AL 1) > 100 × 103 cells/ml or ≥ 20 mg/l biomass (AL 2)

MCs

Minimum Frequency

Maximum limit

France

Cyanobacteria

Toxin legislatedRecommendation

Country

Tab. 13.3 (continued)

450 | Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana

MC-LR

MC-LR

No federal regulation. State specific regulations

MCs

MCs

MCs CYL ANAT-a STX

Uruguay

South Africa

USA*

Florida

Ohio

Oregon

1–12 μg/l 1 μg/l 3 μg/l 3 μg/l

1 μg/l

1 μg/l

1 μg/l

1 μg/l

Eq.: equivalents, Resp.: respectively, SL: surveillance level.

* New USA legislation on cyanotoxins is expected in 12–24 months from the publication of this book.

From Chorus 2012 [49] and Ibelings et al. 2014 [47].

Biovol.: biovolume,

1 μg/l MC-LR eq. > 5 × 103 cells/ml > 1 μg/l chlorophyll-a.

MCs Cyanobacteria

Turkey (proposed legislation)

AL: alert level,

1 μg/l

MCs

Spain

1 μg/l

MC-LR

Singapore

Maximum limit

Toxin legislatedRecommendation

Country

Tab. 13.3 (continued)

Cyanobacterial populations monthly

Only under suspicion of eutrophization

Annual

Minimum Frequency 13 From science to policy: dynamic adaptation of legal regulations | 451

Cells Chlorophyll-a Scum

Cells Biovol.

WHO

Australia

100 x 103

Cells Chlorophyll-a

Visual inspection Chlorophyll-a Microscopy Toxin

Czech Republic

Denmark

Presence of scums > 50 μg/l Cyanobacteria dominate Some regions

Visual inspection routinely

Visual inspection monthly Sampling four times a year

< 1.5 × 103 cells/ml (SL) < 0.5 × 103 cells/ml (SL) 20–100 × 103 cells/ml and > 50 % cyanobacteria (AL 1) Toxic cyanobaceteria sp. (AL 1) Humans or animals (AL 2) Consistently (AL 2)

Phytoplankton Cyanobacteria Phytoplankton and cyanobacteria fraction Report of toxic effects Scum presence

Cuba (no regulations, framework being tested)

20 × 103 or 100 × 103 cells/ml for AL 1 or 2 resp.)

Variable among regions

20 μg/l 100 × 103 cells/ml

Annual (risk assessment)

Minimum Frequency

MC-LR Cells

< 5 × 103 cells M. aeruginosa/ml or < 0.4 mm3 /l biovol. for SL) < 50 × 103 cells M. aeruginosa/ml or < 4 mm3 /l biovol., < 10 mm3 /l biovol. if no known toxin producer present (AL 1) > 50 × 103 cells M. aeruginosa/ml, or > 4 mm3 /l biovol., > 10 mm3 biovol. if no known toxin producer present (AL 2) 10 μg/l MCs (AL2)

20 x or cells/ml for AL 1 or 2 resp.) 10 or 50 μg/l chlorophyll-a. (AL 1 or 2 resp.) Visualized

103

Limits

Canada

MCs

Parameter regulated

Country

Recreational water

Tab. 13.4: Legislation on cyanotoxins’ presence in recreational water.

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Parameter regulated

Visual inspection

Visual inspection Cyanobacteria MCs Scums

Transparency Biovol. Chlorophyll-a MCs Scums

Cells Chlorophyll-a MC-LR eq.

Cells MCs Scums

Chlorophyll-a Cyanobacteria Scum

Country

Finland

France

Germany

Hungary

Italy

Netherland

Tab. 13.4 (continued)

Daily observation. Weakly counting Fortnightly

< 12.5, 12.5–75 or > 75 μg/l cyano-chlorophyll-a. (SL, AL 1 or 2 resp.) < 2.5, 2.5–15 or > 15 mm3 /l biovol. (SL, AL 1 or 2 resp.) Intensities I, II and III (SL, AL 1 or AL 2 resp.)

Daily

Daily

Visual inspection weekly

Minimum Frequency

< 20, 20–100 or > 100 × 103 cells/ml (SL, AL 1 or AL 2 resp.) > 25 μg/l (AL 2) Presence (AL 2)

< 20, < 50, < 100 or > 100 × 103 cells/ml (SL 1, SL 2, AL 1 or AL 2 resp.) < 10 , < 25, < 50 or > 50 μg/l chlorophyll-a. (SL 1, SL 2, AL 1 or 2 resp.) < 4 , < 10, < 20 or > 20 μg/l (SL 1, SL 2, AL 1 or 2 resp.)

Secchi Disk reading < 1 or > 1 m for AL 1 or 2 resp. < 1 or > 1 mm3 /l biovol. (AL 1 or 2 resp.) < 40, > 40 μg/l chlorophyll-a. (AL 1 or 2 resp.) < 10, > 10 or > 100 μg/l MCs for AL 1, 2 or 3 resp. AL 3

Bloom < 20 × 103 , 20–100 × 103 or > 100 × 103 cells/ml for AL 1, 2 or 3 resp. 25 μg/l MC-LR eq. for AL 4 AL 4

Transparency, greenish flakes, colored water, heavy surface scums (SL, AL 1, 2 and 3 resp.)

Limits

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Cyanobacteria

New Zealand

MCs Cyanobacteria Scums

No federal regulation. State specific regulations in 21 states

Turkey (proposed legislation)

USA

From Chorus 2012 [49].

Eq.: equivalents, Resp.: respectively, SL: surveillance level.

Fortnightly

< 10, 10–25 or > 25 μg/l MC-LR eq. (AL 1, 2 and 3 resp.) < 20 or 20–100 × 103 cells/ml (AL 1 or 2 resp.) AL 3

Probability of cyanobacterial proliferation

Spain

AL: alert level, Biovol.: biovolume,

Variable depending of risk estimation

Low, medium, high probability classification per site

Chlorophyll-a

Singapore

Annual

National implementation EU directive ≤ 50 μg/l for 95 % of three year period

Fortnightly or weekly if area susceptible of cyanobacteria blooms

< 500 cells/ml (SL) 0.5–10 or > 10 mm3 /l biovol. (AL 1 or 2 resp.) 0.5–1.8 or > 1.8 mm3 /l biovol. of toxic sp. (AL 1 or 2 resp.) 12 μg/l MC-LR eq. (AL 2) Consistent presence (AL 2) < 20, 20–50 or > 50 % of surface (SL, AL 1 or 2 resp.)

four times per season

Minimum Frequency

Limits

Poland

MCs Scums Benthic mats

Parameter regulated

Country

Tab. 13.4 (continued)

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13 From science to policy: dynamic adaptation of legal regulations | 455

Tab. 13.5: Legislation on cyanotoxins’ presence in freshwater food. Freshwater food Country

Food sample

Limits

Canada

No national regulation

Victoria (health alert levels)

Whole organism (wet weight)

24–51 μg/kg MC-LR or similar hepatotoxins 18–39 μg/kg CYN and deoxy-CYN 800 μg/kg STX

France (acknowledgement of risk)

Fish

5.6 μg/kg for adults 1.4 μg/kg for children

USA

No federal regulation. State specific.

California

Fish (wet weight)

10 μg/kg MC (LA, LR, RR, YR) 66 μg/kg CYN 1.1 mg/kg ANAT-a

From Ibelings et al. 2014 [47].

13.4 New occurrences of toxic episodes challenge protection of consumer’s safety There have been multiple appearances of new toxins or disseminations of certain toxins to new locations in the last 15 years. Some examples will be discussed in this section because they have escaped and challenged consumer safety protection programs. One of the most remarkable new occurrences was the appearance of a completely new class of toxins, the AZAs, that caused the first human poisoning episode in 1995 in Europe, followed by other outbreaks in 1997, 1998, 2000 and 2008 [8, 52, 53]. Consequently, the presence of AZAs in shellfish was regulated in Europe in 2004 [20]. Although initially AZAs seemed to be exclusively distributed along European coasts [54–58], new locations have been recently described in Chile, the northwestern coast of Africa, Mexico and US [59–64]. The spread of AZAs to a worldwide distribution has prompted international guidelines on their regulation and legal action limits have been recently established in several countries (Tab. 13.2); for example, in the USA, action limits for AZAs were regulated in 2011 [15]. Another increasing toxicological threat is related to ciguatera. Several outbreaks have been reported in areas not considered endemic of ciguatoxic fish, such as New York, North Carolina or Germany [65–67]. These cases were due to the import of reef fish species from tropical or subtropical areas where ciguatera is endemic. Clearly, the increase of international fish trade is an important determinant of ciguatera epidemiology [68]. However, the frequency of ciguatera episodes and ciguatoxic fish captures seems to be increasing in tropical/subtropical areas, such as the western Gulf of Mexico, Puerto Rico and Pacific Islands [69–71]. In 2004, ciguatoxic fish captured in

456 | Natalia Vilariño, M. Carmen Louzao, María Fraga, and Luis M. Botana

the African coasts of the Canary Islands (Spain) caused ciguatera episodes in this location for the first time. This area was not considered endemic for this disease [72]. Later, other human intoxications from fish captured in African coasts and the eastern Mediterranean Sea have been reported as ciguatera, although the presence of Caribbean or Pacific CTXs was not confirmed by analytical methods [73, 74]. In 2008, CTXs were also detected in fish samples captured in Cameroon and Madeira [75, 76], and additional outbreaks of ciguatera fish poisoning in the Canary Islands were reported [77], supporting the suspicion that ciguatera might be a growing problem in the northwestern coast of Africa. Diarrheic shellfish toxins (DSPs) have been monitored for years in multiple locations around the world such as Japan and Europe; however, their occurrence had never been reported in the USA’s waters until 2008 when shellfish contaminated with OA in Texas forced the closure of the affected harvesting area [78]. Therefore, DSPs were not included in routine monitoring programs for marine toxins in the USA until 2011 when the appearance of DSP toxins in Washington State caused a human poisoning outbreak [64]. This episode prompted the inclusion of DSPs in routine monitoring of shellfish toxicity in Washington State. TTX has been typically related to finfish intoxication due to consumption of puffer fish or “fugu”, a Japanese delicacy, captured in the Indo-Pacific Ocean regions. Toxic puffer fish species have also been described in other areas of the globe, such as Florida, Mexico or Greece [79–81]. Their toxicity may be owed to the presence of TTX but also to the presence of PSP toxins (STX group) [80, 82]. An increase of accidental captures of puffer fish in recent years has also occurred along the northern coast of Spain, although toxicity of these specimens has not been reported yet. Interestingly, in 2007 an acute intoxication appeared in Spain due to consumption of Charonia lampas lampas contaminated with TTX [83–85]. This gastropod specimen had been captured in Portugal and in later studies an extended presence of TTX in Charonia lampas and other gastropod vectors along the coast of Portugal was confirmed [86]. Although TTX is known to be present in many marine, freshwater and terrestrial species, and it has been linked to human intoxication by ingestion of gastropods before [87], its occurrence in mollusks along European coasts had not been previously described. In addition, an increase of Tetradontiformes numbers in European Atlantic Ocean has been reported [88], and the recent appearance of the tropical/subtropical puffer fish Lagocephalus laevigatus along the northwest coast of Spain constitutes the northernmost occurrence of this toxic species [89, 90]. Lately, TTX has also been detected in bivalve mollusks harvested from the south coast of England for the first time [91]. These data clearly demonstrate a potentially increasing toxicological problem along European coasts that is not adequately regulated by current legislations or monitoring programs. It could be argued that the presence of TTX is regulated in shellfish because it causes a paralytic poisoning syndrome, and it could be classified as a PSP toxin. However, TTX is chemically different from STXs and it is currently not being detected by the analytical methods used for the detection of PSPs. While the use of the mouse

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bioassay for PSP detection may also yield positive results when TTX is present, the substitution of bioassays for analytical methods do not warrant protection against TTX. Moreover, although regulations may refer in general to shellfish, most monitoring programs are focused on toxin detection in bivalve mollusks, not gastropod mollusks. Palytoxins (PLTX) can also cause human poisoning through ingestion of contaminated seafood, but the number of reports is lower than for other toxins. However, reports of PLTX-like intoxications in Japan and the presence of PLTX-like molecules in shellfish in Europe may be indicators of an upcoming problem [6, 21]. Recently, a new route of exposure to PLTX-like compounds threatening human health has been suggested after the appearance of an Ostreopsis ovata bloom in seaside regions of Italy in 2005 [92, 93]. Inhalation of marine aerosols during this unusual bloom was related to the appearance of respiratory problems, fever and conjunctivitis in humans, in some cases requiring hospitalization. O. ovata from this and posterior blooms was later described to produce several PLTX-like compounds [92–95]. The toxicology through this route should be evaluated for the confirmation of a causal relationship between PLTXs and respiratory distress and other associated symptoms. If confirmed, provisions should be made to regulate the use of marine recreational waters or certain beach locations similar to the process for freshwater recreational waters. Finally, in 2008 PLTX poisoning through dermal exposure by contact with aquarium zoanthids was also confirmed [96] after numerous informal reports from aquarium hobbyists about numbness of hands and arms that they related to the handling of zoanthids [93]. Several clinical case reports followed [97, 98]. Warning and handling instructions on certain aquarium species should also be regulated at the very least.

13.5 Limitations for the development and implementation of new regulations: from science to policy or from policy to science? This scenario of marine and freshwater toxin dissemination or globalization due to new appearances or international trade requires new regulatory measures to protect human health. However, there are still important limitations that preclude the establishment of adequate action limits for some toxins and the implementation of the regulated levels. There is a clear need to improve toxin detection and toxicological characterization of several toxin groups.

13.5.1 Technical limitations for recent/future toxin regulations From a practical perspective, the establishment of regulations that cannot be implemented is useless. At present, there are several technical issues that limit implemen-

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tation of current regulations and also limit prescription of new regulations. These limitations are mainly related to the lack of toxin standards and detection techniques. An urgent issue to be solved in the field of aquatic toxin detection is the scarce availability of commercial certified reference standards for most of the compounds that belong to these toxin classes [4, 6–9, 11, 12]. The currently available toxin standards are listed in Tab. 13.6. The need for standards has become critical with the replacement of biological methods by analytical methods. Analytical methods require calibration with each analog to allow identification and quantification [99, 100]. In addition, this calibration should be performed with certified reference material, which is only provided by a few commercial sources such as the National Research Council, Canada and Laboratorio CIFGA, Spain. Unfortunately, for most groups there are only certified reference materials for a few congeners. In some cases, including CTXs, PLTXs and BTXs, no certified standards are commercially available. With regards to detection techniques, several problems need to be addressed in the near future. Analytical techniques, mainly LC-MS/MS-based methods, are progressively replacing biological methods as a reference or confirmatory method to determine biotoxins. Although these techniques are highly specific, quantitative and allow multiple toxin detection, processing of a high number of samples through these methods will be difficult to achieve for routine testing laboratories [99]. A huge economic effort will be necessary to warrant fast delivery of LC-MS/MS results in dynamic monitoring programs. Therefore, high through-put methods should be developed for rapid, preliminary screening of high numbers of samples in order to reduce the samples to be processed by expensive, confirmatory analytical techniques [99]. The validation of more cost-effective analytical methods for lipophilic toxins would also help to solve this problem [101]. The requirements of chemical methods for the analysis of SXT, OA, DA and AZA groups are defined in the CODEX STAN 292-2008 document revised in 2014 [14]. The criteria to be met are LOD and LOQ or minimum applicable range, precision and recovery. Values for these parameters are specified for several toxin analogs of each group. Methods will have to be validated for these relevant analogs according to validation guidelines and regulations [102, 103]. No minimum performance criteria have been published for CTXs, TTXs, BTXs, cyanotoxins or other toxin groups, despite the fact that provisions on BTXs have been included in the CODEX STAN 292-2008 [14] and BTXs and the other groups are regulated in many countries. Actually, action limits have been regulated recently for CTXs and cyanotoxins despite the lack of adequately validated methods for their detection. The situation is more complicated for non-analytical methods, which in many cases could be useful as screening tools [100]. No criteria or guidelines for the minimum requirements that these methods should fulfill are available. Many nonanalytical methods have been developed in the last two decades for the detection of marine toxins (for a review, see [100, 104]) and only two of them have become official methods, the receptor-based assay for PSP toxins [25] and an ELISA for DA [38, 39].

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Tab. 13.6: Currently available aquatic toxin standards. Toxin group

Toxin

Standard availability*

PSP toxins

STX dcSTX GTX1, 4 GTX2, 3 dcGTX2, 3 GTX5 C1, 2 NEO dcNEO

CRM CRM CRM CRM CRM CRM CRM CRM CRM

DSP toxins

OA DTX-1 DTX-2

CRM CRM CRM

Domoic acid

DA

CRM

TTXs

TTX and 4, 9-anhydroTTX

CRM

AZAs

AZA-1 AZA-2 AZA-3 AZA-4 AZA-5

CRM CRM CRM QCS QCS

YTXs

YTX homoYTX

CRM CRM

PTXs

PTX2

CRM

Cyclic imines

20-methyl SPX G 13-desmethyl SPX C 13, 19-didesmethyl SPX C Gymnodimine

CRM CRM QCS CRM

Microcystins

MC-LR MC-RR MC-YR MC-LW MC-LF MC-LY [D-asp3 , E-Dhb7 ]-MC-RR dmMC-LR Nodularin-R

CRM CRM AS AS AS AS AS CRM CRM

Cylindrospermopsins

CYN

CRM

Anatoxin-a

ANAT-a

CRM

Matrixes

DA-contaminated mussel AZA-contaminated mussel DSP-contaminated mussel Non-contaminated mussel PSP-contaminated Pacific oyster

CRM CRM CRM CRM CRM

* Only highest quality available is reported. CRM: certified reference material, QCS: quality controlled standards, AS: analytical standard.

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One of the drawbacks of non-analytical methods is that most of the developed assays have not been appropriately validated. The high diversity of these methods makes it difficult to establish general guidelines for their performance, and the absence of guidelines discourages validation efforts. The detection of CTXs for efficient protection of human consumers will require a different approach to toxin detection in shellfish. A validated mouse bioassay has been published for CTX detection in fish [105, 106]; however, it is not used for toxin detection before commercialization [107], only for epidemiological studies or postconsumption confirmation of ciguatera [16, 107, 108]. In addition, its sensitivity for P-CTX-1 and C-CTX-1, with LODs of 0.2 and 3 μg/kg respectively [6], is not enough to comply with current regulatory limits. Other epidemiological studies are based on the use of vitro bioassays for screening and analytical methods for confirmation and profiling [109, 110]. Chemical determination of CTXs is not an easy task owed to the high number of analogs, the complexity of these molecules, the lack of standards and the low sensitivity for this class of toxins [74], and therefore routine testing has not been implemented in most monitoring laboratories [109]. In recent years, the sensitivity of analytical methods, mainly HPLC coupled to mass spectrometry detection, has improved with new technological developments reaching LOQs from 5–100 ng/kg [74, 111, 112]. However, analytical detection of CTXs still cannot reach the sensitivity necessary for legislated action limits, considering that although LOQs are close to regulatory levels, they refer to individual analogs and in reality total sample toxicity is the result of a mixture of CTXs [68, 112, 113]. Because the action limits are legislated in CTX equivalents and therefore all congeners should be considered in calculating final CTX equivalents, individual LOQs will have to be considerably lower than the action limit. Moreover, sample preparation is too laborious for routine testing of every suspected specimen [15, 74, 111, 112]. Some non-analytical methods, such as the cytotoxicity assay and the receptor-based assay, provide high sensitivity with LOQs around 39 ng/kg and 155 ng/kg respectively [114, 115], but it is clearly not enough to warrant full implementation of current regulations. A more sensitive cell-based assay that uses enhancement of response by veratridine provides LOQs in the pg/kg range [6, 116]. In any case, none of these methods is adequate for efficient pre-market testing, even if only known ciguatera vectors are tested. Fish are not sessile organisms, and therefore, detection would have to be done on a per unit basis instead of on a whole batch [107]. These characteristics demand a fast, low-cost, simple and reliable tests to warrant protection of consumers. In addition, the sensitivity of the assay will have to be extremely high to comply with current action limits recently regulated in some countries [15]. Although several non-analytical methods have been developed [74, 104], none of them gathers all the characteristics necessary for routine CTX testing. Two immunoassays were commercialized for CTX screening, but their production has been ceased, probably because they did not seem to provide reliable performance [74, 107].

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Similarly, BTX detection has been historically done by MBA, although analytical and other non-analytical methods are available [4]. Analytical methods are often used for epidemiological studies and confirmation purposes. Non-analytical methods such as ELISA and other immunoassays are commercially available for sample screening. In this case, most analytical and non-analytical methods have enough sensitivity to comply with current action limits for BTX. However, inter-laboratory validation has not been done for most of them and none has become an official method. For freshwater cyanotoxins, there is no regulation on detection methods or their minimum performance requirements. Many analytical and non-analytical methods have been published for cyanotoxins, with estimated sensitivity that would allow detection at the regulated limits [100], but adequate inter-laboratory validation has not been done for most of them [100]. Another task that will have to be addressed for suitable implementation of regulations is validation of detection methods in the routine testing laboratory. Validation should be performed for every species suspected of contamination [102, 103]. For most toxin groups an important number of edible shellfish or finfish species have been reported as vectors for human intoxication. For example, more than 50 finfish species have been described to contain CTXs (see Chapter 9). Besides the huge effort to validate a detection method for several species, some species may be specially challenging due to the presence of matrix effects. Finally, there are some issues that cannot be solved with technical development. The EU legislation indicates that CTXs are not to be present in seafood [20]; however, the absence of toxin cannot be demonstrated by any detection method. No method can detect amounts below its LOD. This limitation can only be solved by changing the text of the regulation and establishing a maximum permitted level. The same situation applies to TTX or paralytic toxins in fish [20].

13.5.2 Toxicological limitations for new toxin regulations One of the most important limitations for new regulations is the lack of information about toxicological characteristics of the multiple analogs within a toxin class. Usually, regulations specify an amount of toxin for a certain group in terms of equivalents of a representative member of the group. The term “equivalents” refers to the amount of an analog that would cause toxicological signs equivalent to the regulated limit of the representative toxin. Therefore, when using analytical methods for the detection of toxins, it is absolutely necessary to know the relative toxic potency of the analog versus the representative toxin in order to transform the amount of an analog to toxin equivalents. For most aquatic toxin classes, acute toxicity of the representative toxin has been characterized and its potency established in terms of LD50 after i. p. injection. However, for regulated toxin groups, the toxicity of many analogs is unknown due to the scarce amount of pure compounds, which is not enough for toxicological

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studies. In addition, most of the data on toxic potency have not been obtained with certified standards because they are available only for a few toxins (see Tab. 13.6). The relative toxic potency versus the representative toxin is expressed as TEF and should be established for each relevant analog. Unfortunately, except for a few groups, the international scientific community agrees on the lack of experimental data to calculate a TEF value for every analog [4–13]. One of the first official analytical methods accepted worldwide was published for the PSP toxin group in 2006 [24]. Almost ten years later, there are still no definitive international guidelines on the TEF values that should be used for the most common PSP analogs [117]. For this toxin group, there is enough toxicological information to suggest provisional TEFs for the analogs that most often appear in contaminated shellfish, although there has been no uniformity in the values used by different laboratories [41]. Recently, internationally agreed TEFs have been assigned to 17 PSP analogs of the more than 50 naturally occurring compounds described for this group [118]; they are available in the Codex web site [41]. It is expected that guideline TEFs will be officially included in the “Standard for live and raw bivalve molluscs” of the Codex Alimentarius in the future [14]. The situation is similar for the DSP and AZA groups. DSP toxins comprise OA, DTX-1 and DTX-2 and their acyl-esters, and TEFs have been proposed for the three compounds. In the AZA group, more than 20 naturally occurring analogs of AZAs have been reported, however three of them are the most prevalent and therefore in many countries only AZA-1, AZA-2 and AZA-3 are regulated [20, 119]. Provisional TEF values for these three AZAs are available in the Codex website as well [41]. The relative toxicity of the other AZAs is not known, except for AZA-4 and AZA-5 i.p. LD50 values, which are less toxic than the 3 regulated AZAs. The presence of these other analogs simultaneously to much higher amounts of AZA-1, 2 or 3 has been the basis for considering that their contribution to toxicity would be irrelevant. However, there is no evidence that this assumption is valid for all the analogs described. The situation is much more complicated for other toxin groups. CTXs have been classified in three different groups based on their origin and chemical structure: P-CTXs (Pacific Ocean), C-CTXs (Caribbean) and I-CTXs (Indian Ocean), comprising 29, 12 and 4 different molecules respectively [113, 120, 121, 123–125]. Intraperitoneal LD50 in mice has been determined for some CTXs present in toxic fish, with values of 0.25–0.35, 2.3 and 0.9 μg/kg for Pacific P-CTX-1, 2 and 3, and 3.6 and 1 μg/kg for Caribean C-CTX-1 and 2 respectively [105, 121, 122, 124, 126]. I-CTXs LD50 was estimated at around 5 μg/kg [120]. The toxicity of the CTXs found in Gambiediscus is lower in general, with i.p. LD50 values of 12, 20, 2.5, 8 and 10 μg/kg for CTX4A, CTX4B, CTX3C, 49-epi-CTX-3C and M-seco-CTX-3C respectively [127]. For other members of the group, values are reported as mouse lethality, without specification of LD50 or MLD (minimum lethal dose) [128], and therefore they should not be considered adequate for TEF value calculation. Obviously, no certified standards were used in these studies. P-CTX-1 is considered to dominate the toxicological profile of ciguatoxic carnivorous reef fish in the Pacific [121, 122], although other analogs may be responsible for more

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than 25 % of toxicity in some samples [112, 129]. In fish from the Caribean or Atlantic Ocean, it does not seem to be a clear dominance of one or two Caribbean CTXs [68]. Therefore, the lack of information on toxic potency of P-CTXs might not be so critical assuming a prevalence of P-CTX-1, but this is not the case for C-CTXs, where C-CTX-1 and 2 have been reported to contribute only to 40–60 % of total CTX-related toxicity [123]. The studies providing information about the toxic profile of ciguatoxic fish are very scarce. The presence of CTXs has been reported in more than 50 species, and the CTX profile has only been described in a few of them (see Chapter 9); therefore, more studies are necessary to clarify the prevalence of CTX analogs in order to determine their relevance for regulation purposes. In addition, more toxicological data should be obtained in order to establish TEF values for the different CTXs. The EFSA CONTAM panel on CTXs adopted TEF values for several CTX analogs based on i.p. LD50 [6], although, as discussed above, some of the data used for this calculation are not clearly reported in the literature as LD50 [128]. Interestingly, USA and Mexican legislations regulate the amount of CTXs in equivalents of P-CTX-1 or C-CTX-1 [15, 18], but currently there is no official information about internationally agreed TEF values that should be used if samples are detected with analytical methods. In most countries, the regulations on TTX are limited to the prohibition of the import and sale of puffer fish. In the EU, the regulation of an absence of toxin also applies to TTX, and we have already discussed that the absence of toxin cannot be demonstrated. Recent occurrences of TTX in mollusks in Europe have raised concerns about consumers’ safety, which may in the future lead to the regulation of a maximum content of TTX in seafood. At least 15 naturally occurring members of the TTX group have been described in marine organisms [130–132]. LD50 data can be found in the literature for some of them, and, at present, TTX is the most toxic compound of the group. The i.p. LD50 of TTX is estimated in 10 μg/kg; for 11-deoxyTTX, 6, 11-dideoxyTTX and 11-norTTX-6(S)-ol the estimation is 70, 420 and 54 μg/kg; 4, 9-anhydroTTX, 5-deoxyTTX, 5, 6, 11-trideoxyTTX and 4-S-cysteinyl-TTX are considered almost non toxic [82, 131–136]. TTX seems to be the more abundant analog in seafood, but other analogs considered to be non-toxic have also been reported in similar amounts in some samples [82]. Other toxic analogs, although less toxic than TTX, can also contribute to overall toxicity. As an example, 4-epi-TTX, which is considered to be six times less toxic than TTX, and 11-norTTX-6(S)-ol, which would be five times less toxic than TTX, are present in a relevant amount in some samples [82, 130–132, 135, 137, 138]. In addition, the profile of TTX analogs varies among toxic animal species and organs [82, 139]. Further studies will be necessary to propose TEF values for the compounds of this group. Toxicity was not determined in homogeneous conditions; sometimes the experimental methodology is not even reported, no standards were used for these studies and on occasion the LD50 values are reported as “roughly determined” [133]. A change in TTX regulation from a ban on puffer fish species commercialization to establishment of a maximum limit of TTXs in seafood will require data on the occurrence of different analogs in the regulated

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species and toxicity of those relevant analogs in order to make an appropriate risk assessment. BTX-related toxicity evaluation by analytical methods faces the same lack of toxicological data as the other toxin groups. Besides the scarcity of toxicological studies, toxicity is reported as LD50 for a few analogs, while for other congeners also present in fish, toxicity is reported as MLD [4]. Therefore, no TEF values have been proposed for BTXs. Likewise, toxicological data on PLTXs is very limited, with only i.p. LD50 values reported for PLTX and osteocin-D of 0.29 and 0.75 μg/kg respectively [140, 141]. Oral toxicity has been studied for PLTX and 42-hydroxy PLTX in mice with LD50s of 767 and 651 μg/kg respectively [142, 143]. At least 17 PLTX-like toxins have been described to date [144]. Microcystins are also a huge group of compounds, with more than 100 analogs [145]. Mouse i.p. LD50 s have been determined for a few microcystins, but the toxicology of most members of the group is unknown. I.p. LD50 s reported for MC-LR, MC-RR and MC-YR are 43–87, 230, and 110.6 μg/kg respectively [146–148]. Intravenous toxicity is also available for some analogs [149], but oral toxic potency is only known for MC-LR, with an LD50 value in mice estimated in 10.9 mg/kg [150]. The provisional limit for microcystins recommended by the WHO refers only to MC-LR. In several countries, however, the regulations are extensive to all the members of the group, although information about their toxicity is lacking. WHO recommendations are not considered adequate nowadays based on occurrence and toxicity of other microcystin congeners [151], but the regulation of “microcystins” in general without specification of analogs and without toxicological data about their toxicity is not appropriate either. The frequency of appearance and the fraction of overall toxicity that can be attributed to an analog must be determined for a high number of samples before establishing reliable data on microcystins’ prevalence and relevant congeners. Therefore, the toxicity of the most prevalent microcystins should also be determined before regulations covering the microcystin group can be implemented. Likewise, cylindrospermopsin and anatoxin-a recommendations are listed in some countries, in spite of the lack of information to make an adequate assessment of the threat that they pose to human health. Another important issue with regards to TEFs is the lack of agreement on the methodology when calculating the final TEF value based on toxicity data. Usually, the ratio of i.p. LD50 s in μg/kg is used instead of mols/kg. The Codex standard document indicates that total toxicity should be estimated “as the sum of the molar concentrations of detected analogs multiplied by the relevant specific TEFs” [14]; consequently, TEFs should be calculated from LD50 s in mols/kg. Finally, it is important to emphasize that most toxicological data have not been obtained with certified standards and therefore the TEFs derived from those data should be considered provisional until confirmed. Furthermore, these toxicological experiments were performed using i.p. administration, which is obviously not the natural route of human exposure to these toxins, and therefore considered inadequate for risk

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assessment and TEF calculation [152, 153]. TEF values should be based on the oral toxicity of these compounds, since toxicokinetics is an important component of overall toxicity. In fact, several PSP analogs showed different toxic potency (relative to STX) when administered orally or i.p. [154]. Unfortunately, the scarcity of pure toxins has precluded toxicological testing by oral administration of most analogs, because this route requires higher amounts of toxin. Oral toxicity is usually known for the representative toxin of the group, but not for the less-frequent or less-abundant analogs [153]. A completely different approach regarding toxicity evaluation will be needed for non-analytical methods. In this case, the assays yield a value of toxin content for the toxin class without individual quantification of analogs [104]. Therefore, TEFs cannot be used to calculate overall toxicity based on the analog amount. For non-analytical methods, the cross-reactivity of the assay for every regulated analog should be determined and it should be equivalent to TEF values [104]. The cross-reactivity values can be highly variable for different assays. Receptor-based and functional assays usually provide cross-reactivity profiles closer to actual toxicity than immunoassays. For example, most immunoassays developed for PSP toxins do not detect neoSTX, one of the most toxic analogs of the group [155, 156]. On the contrary, a reasonable match to toxicity can be found for DSP immunodetection [157]. In vitro functional assays are based on biological interactions or activity of the toxins. In these methods, the kinetics of the toxin in the organism is not accounted for, and actual toxicity may differ from in vitro activity. An example would be the detection of microcystins using the protein phosphatase inhibition assays, which yield highly variable ratios of protein phosphatase inhibitory potencies for the most commonly tested congeners (MC-LR, MC-RR and MC-YR) [158–161]. The relative inhibitory potencies of these MCs are not in agreement with relative i.p. toxic potencies. Therefore, the performance criteria for non-analytical methods should include cross-reactivity requirements for every relevant toxin based on TEF values in the future.

13.5.3 Economic limitations Although most of the technical and toxicological limitations outlined above can be solved, it is needless to say that overcoming these limitations will require a huge economic effort. Research on toxicology and detection methods, with application of new technologies to aquatic toxin detection and/or inter-laboratory validation of existing methods, is – as was pointed out in the previous two sections – urgently needed. Additionally, an important amount of resources will have to be directed to endow routine testing laboratories with instrumentation adequate for toxin detection according to new regulations. This instrumentation should be able to provide results for the number of samples routinely tested in the laboratory in a timely fashion. Training of laboratory personnel will also be required in order to implement new regulations. Instrumentation and maintenance required for HPLC-MS/MS analysis is expensive, and

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impractical for high sample numbers, and from a worldwide perspective, difficult to afford by developing countries [101]. Validation of other analytical methods alternative to HPLC-MS/MS analysis for some toxin groups would help to afford implementation of current and future regulations [101]. A great international effort is being made to achieve uniformity of marine toxin regulations worldwide. Nowadays, harmonization is an impelling demand of seafood market globalization, and a lot of work remains to be done in this endeavor to improve the existing international guidelines and to include other toxin groups. Establishment and implementation of new regulations imply a high cost that society, governments and institutions will have to be willing to cover. If regulations are not accompanied by funding, then implementation and informed/meaningful regulations are not possible.

13.6 Modification of monitoring and surveillance programs Monitoring programs for the protection of human health have been implemented in relation to marine and freshwater toxins. Although there are legal regulations prescribing the controls on bivalve shellfish production to avoid toxin-contaminated shellfish reaching human consumers [162], these regulations outline general rules, and monitoring programs are usually designed considering the local toxicity profile [163]. This approach is efficient to warrant consumer protection while avoiding unnecessary expenses by rationalizing phytoplankton and toxin monitoring (for the design monitoring programs, see [163]). However, several aspects of monitoring programs should be revised periodically and steps for their modification, if necessary, should also be outlined in local marine toxin risk management plans. Several elements of monitoring plans could possibly be affected by climate change, such as identification and quantification of harmful phytoplankton species or the classes of toxins routinely tested. In addition, evolving regulations and technological developments will also influence detection techniques. As a consequence, training of laboratory personnel will also have to be periodically re-evaluated to cope with new species or toxins monitoring and application of new technologies. Some countries have already developed a dynamic monitoring program for cyanotoxins based on the evaluation per site or location of the risks of cyanobacterial blooms annually [49].

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13.7 Integrative example: tetrodotoxin as a biomarker of climate change In past years, there has been an increasing awareness of the actual change in climate due to global warming mostly as a consequence of the frequent extreme weather episodes worldwide that are covered exhaustively by the communications media. But along with the weather,global warming includes many other silent and unnoticed phenomena, such as ecological modifications in the seas and new toxic events that do not receive any attention, and, in fact, are rather difficult to identify by the experts. In this chapter we will use TTX as an example of a new situation that is most probably caused by global warming; it requires attention by food toxicologists but also by legislators. We will try to expose the elements that support a link between global warming and TTX presence in the food chain. In Europe, the Lessepsian migration, also called Erythrean invasion, caused many marine species to enter the Mediterranean Sea from the Indo-Pacific Ocean as a consequence of the opening of the Suez Canal in 1869 (the name comes after the French engineer Ferninand Marie de Lesseps who supervised the construction of the canal). More than 300 species from the Red Sea were identified in the Mediterranean. This man-made modification of ecology is not related to climate. However, since the Mediterraneum is connected to the Atlantic, it allows for the identification of a clear progress of how a species, otherwise not present before, suffers the impact of climate change after they have settled in their new niche. At least 67 species of fish were identified as Lessepsian immigrants, many of them poisonous, with species of the families Siganidae, Dasyatidae, Scorpanidae, Plotosidae, Ostracionidae and Ariidae [164, 165]. Clinical intoxications were reported by ingestion of some of these fish species, as some of them are known to have CTXs [166, 167] or TTX [168]. In the case of TTX, the presence of the Lessepsian inmigrant fish Lagocephalus sceleratus (Gmelin, 1789) is well reported on the eastern Mediterranean Sea, along the coasts of Israel, Palestine and Libanon [168] and in the Aegean Sea by the coast of Greece [79]. It is known as silverstripe blaasop, and it is a pufferfish of the family Tetraodontidae. The first report of this fish in the Mediterranean Sea was in 2003, and now it is a common presence in the waters and considered the fastest expanding Lesseepsian immigrant. Since the European regulation [20, 162] states that fish of the family Tetraodontidae are not to be placed on the market, specific measures are being taken to avoid this, especially when these fish are collected along with commercial species. Therefore, the presence of these toxic fish increases press coverage and fishermen’s awareness [79]. Recently, in the Cantabric Sea, by the North coast of Spain, the collection of several individuals of Lagocephalus spp. in 2014 was also covered by the press, as this was further north than the presence previously reported in Galicia, northwest Spain, in 2011 [89].

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All these man-made changes are not related to climate change, although the spreading of TTX-bearing species to the north may be related to seawater warming. Local factors related to climate change are further amplified by anthropogenic effects [169], and the term “climate change” includes both anthropogenic and largescale changes in oceanic patterns. Three decades of high-resolution study of sea surface temperatures show that coastlines are warming up, also with a significant decrease of cold events [170]. In addition to this, although typical on the Iberian peninsula, in northern Europe the presence of anchovy and sardine species is rare, but there is an increase in the abundance of both anchovies and sardines in the northwestern North Sea since 1995 [171], and this corresponds to an ecosystem change occurring in the North Sea that affects the whole food chain due to rising temperature and climate change. A recent survey between 1965 and 2012 again show that warming shelf seas drive subtropicalization [101, 172] of European pelagic fish [171]. It is estimated that Northeast Atlantic mean sea temperature has increased by 1.31 °C in the past three decades [173], the North Sea being one of the global warming key spots. Along with shifts from certain fish species to the north due to sea warming, phytoplankton experienced similar northward distribution changes [174], although it is unclear how this is specifically affecting toxic phytoplankton based on a ten-year study carried out along the coasts of Holland [175]. However, since 50 % of global primary productivity is contributed by phytoplankton, it is reasonable to assume that toxic phytoplankton and the whole ecosystem will be largely affected by climate change. This possibility is discussed by Hallegraeff [176], showing that global distribution of some toxic species, such as Pyrodinium bahamense in recent plankton is reduced when compared to fossil cyst records that show a wider distribution (see also Chapter 6). Also, the climate El Niño event has been linked to massive toxic PSP episodes of Alexandrium tamarense in the Northwest Pacific and of Karenia dititata in Hong Kong [176]. The CTX producer Gambierdiscus spp. has been reported to expand its geographical reach in the past years, from well-known tropical areas to the Mediterranean Sea [177], Canary Islands [72] and Madeira [76]. In fact, it has been projected that an increase in sea temperature may trigger the incidence of ciguatera in Papua New Guinea from 35–70 per thousand people reported in 1990 to 160–430 per thousand by 2050 [178], a trend that may be expected in other parts of the world [179]. Fossil records show that Gymnodinium was very abundant in warm periods (2000 to 500 B. P. (before present)) to almost disappearing during a cool period after 300 (B. P.) [180]. A similar trend in terms of a shift toward Atlantic waters has been observed with the presence of ostreocins from Ostreopsis spp. in Mediterranean coastal waters. PLTXs, ovatoxins, ostreotoxins, mascarenotoxins and ostreocins are very potent and toxic molecules produced, with different chemical profiles, by Ostreopsis spp. [181]. The mechanism of action of PLTX is the inhibition of the Na+ -K+ ATPase, thus altering the equilibrium of sodium and potassium [182]. In 2006, Ostreopsis ovata algal blooms were reported to be the cause of human intoxication in beach tourists

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after breathing wave aerosols in Genova, Italy [183] and the Mediterranean French coast [184]. Later on, the presence of PLTX-like compounds was reported in 2008 in mussels collected from 2004 from the Aegean Sea in Greece [177], and confirmed by a sensitive neuroblastoma-based cytotoxicity assay [185]. It soon became evident that Ostreopsis spp. blooms needed some sort of risk management, as ovatoxins were identified in mussels from Italy, from France and elsewhere in the Caribbean and New Zealand [7, 186]. Further evidence of the presence of Ostreopsis cf. ovata in Spain was demonstrated in 2015 [187], clearly showing a trend along the European Mediterranean coast. The problem of TTX is different than that of PLTX-like compounds, as TTX is produced by bacteria, while PLTX is produced by dinoflagellates. TTX is a thermostable compound originally isolated from the ovaries of pufferfish Fugu rubripes [188], and its structure was independently reported in several laboratories [189–191]. TTX is present in all tissues of pufferfish, with different levels depending on the tissue and the species [139]. It is a site 1 sodium channel blocker, with a mechanism of action similar to STX, although the so-called TTX-sensitive channels are different than those blocked by STX. In particular, TTX is very potent on the subtypes Nav 1.1, Nav 1.2, Nav 1.3 and Nav 1.7 [192], while STX is especially potent on the subtypes Nav 1.2 and Nav 1.6 [193]. The sharpest difference between STX and TTX resides in their effect on the Nav 1.7 subtype, typical of pain sensory neurons. Since it has caused human intoxications and deaths, the toxic levels are well understood, with lethality to humans at about 2 mg [87] or 10 000 mouse units (a MU is the dose that kills a 20 g mice in 30 min, by intraperitoneal administration, reported to be 0.178 μg [194]). The symptoms of the intoxication are similar to the paralytic shellfish toxin intoxication (STX and analogs), namely tongue and lip numbness, facial paresthesia, sense of lightheadedness and floating, unconsciousness and respiratory failure. It is therefore a fact that global warming is causing a northward shift of species. But this also affects the bacterial population. The microorganisms that produce TTX are symbiotic microbes common to many animal species that cover six different phylum [195]. It is unclear why TTX is so common in nature, but in 1986, TTX-producing organisms isolated from xanthid crab showed that Vibrio spp. was a source of TTX, demonstrating that TTX was an exogenous compound [196]. So far 23 bacterial genera have been reported to produce TTX (Vibrio spp., Pseudomonas spp., Serratia marescens, Raoultella terrigena, Roseobacter sp., Aeromonas spp., Shewanella spp., Marinomonas spp., Plesiomonas sp., Alteromonas sp., Acitenobacter sp., Caulobacter sp., Microbacterium sp., Micrococcus sp., Kytococcus sp., Cellulomonas sp., Actinomycete sp., Nocardiopsis sp., Streptomyces sp., Bacillus spp., Lysinibacillus sp., Flavobacterium sp., and Tenacibaculum sp.) [194], of which Vibrio spp. is the main producer. This is important for two reasons,: one is that this production of TTX causes the accumulation of the toxin through the food chain; the second is that global warming is spreading some of the microorganisms northward.

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European legislation [20] for marine toxins applies to bivalves, gastropods, echinoderms and tunicates. In 2008, a serious human intoxication was reported in Spain after consumption of a gastropod, the trumpet shell Charonia lampas lampas [84, 85]. Although this [21] was new for Europe, in 1981, it was reported that TTX was present in trumpet shells Charonia sauliae [197] in Japan, and intoxication from ingestion of gastropods has been reported there on several occasions [87]. This human intoxication has stressed the fact that European legislation needs to cope with current toxicological findings, since as it is written now, the legislation is limited to prohibiting fishing individuals of the Tetraodontidae, Molidae, Diodontidae and Canthigasteridae families [22], hence it does not protect consumers from this potential risk. If new legislation is issued, it would require defining the method of analysis or detection, the limits and the compounds to be monitored. The compounds identified in the Spanish intoxication were mainly TTX (31.5 mg/100 g digestive gland flesh) and a threefold concentration of 5, 6, 11-trideoxyTTX (100.4 mg/100 g), but the chemical diversity of TTX analogs covers many more compounds [139]. In a thorough sampling study carried out after this intoxication, it was found that TTX is a common presence in several gastropod vectors in Portugal, specially Gibbula umbilicalis and Monodonta lineata [86]. The measured amount of TTX and the analog 4-epiTTX was 90.5 and 21.48 ng/g, respectively. A recent report has further evidenced the need for updating legislation: TTX has been reported to be accumulating in mussels (Mytilus edulis) and oysters (Crassosteas gigas) collected in the south of England in 2013 and 2014 [91]. The authors report low but frequent detectable levels (a maximum of 0, 12 mg/kg) of TTX, 5, 6, 11-trideoxy TTX and 4, 9-anhydro TTX, which they traced back to Vibrio parahemolyticus in most of the cultures and one case of Vibrio cholerae. In the work, the authors stress the fact that global warming is related to a higher incidence of Vibrio spp. in the English coasts along the Channel.

13.8 Concluding remarks Global warming seems to be one of the factors involved in the increased distribution of toxic episodes. Consequently, many countries will have to modify their regulations and monitoring programs to warrant protection of human health upon the appearance of new toxicological threats along their coastlines. A remarkable effort is being made to establish international guidelines for the harmonization of aquatic toxin regulations and minimum requirements of detection methods; this will favor quick adaptations to new regulatory needs. In spite of the huge progress in recent years, there are still limitations for determining and implementing new and existing regulations. Toxicological data are scarce for most toxin groups and preclude adequate risk assessment and TEF value estimations. In addition, the validation of fast, cost-effective, confirmatory methods following current international guidelines is an urgent requirement, as is the production of

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fast, reliable and easy-to-use screening tests for toxin detection in the field, at least for some toxin groups. Furthermore, one of the main problems of aquatic toxin detection is the lack of commercial certified standards, which are absolutely necessary for the validation of analytical methods. More changes are expected to occur with progressive global warming and therefore monitoring programs should be dynamic, with general management plans that include periodic revision of monitoring adequacy to the ongoing toxicological situation in one area, but also of suitability to protect against possible threats in the near future. Appropriate protection of human health should be supported by policy and science, and coordinated efforts of regulatory agencies and scientists are necessary for the transfer of knowledge to develop sensible regulatory measures.

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Index Acute reference dose (ARfD) 430, 441, 442 – azaspiracid 442 – berevetoxin 442 – ciguatoxin 442 – domoic acid 442 – okadaic acid 442 – palytoxin 442 – pectenotoxin 442 – saxitoxin 442 – yessotoxin 442 Alexandrium 273 Alexandrium tamarense species complex 128 Alexandrium tamarense/catenella species complex 185 Algal bloom see Cyanobacteria, 196 Algal bloom range extension 182 Algal-bacterial mutualism 324 Alien species 243 – emerging risk 243 – TTX 243 Allelopathic chemical 316 Allelopathic control 335 Allelopathy 134, 335 Amnesic shellfish poisoning 428 Anabaena 206 Analytical method 428 – limitations 458 Anatoxin 156, 167, 427 Anatoxin-a 157 Anatoxin-a(S) 158 Animal 419 – cultured animal 421 – domestic animal 421 – marine animal 419 Antarctic 4 – Antarctic Circumpolar Current 23 – Antarctic Circumpolar Wave 12 Antarctica 168 Anterio-posteriorly compressed 274 Antigen-ovalbumin (A-OVA) 60 Aphanizomenon 206 Aptamer 75 Aquaculture 419 – aquaculture species 419 Arctic 4, 168

Atomic cluster 77 Azaspiracid – occurrence 455 Azaspiracid shellfish poisoning 428 Barley straw 335 Biodegradation 329, 330 Biological oxygen demand (BOD) 323 – BOD 323 Biomanipulation 326 Bioreceptor 47 Biosensor 36 Biosynthesis of STX 131 Brevetoxin 257, 427 – Karenia brevis 258 – Neurotoxic Shellfish Poisoning (NSP) 257 BTX-group toxin 258 – Chattonella antiqua 258 – Chattonella marina 258 – Fibrocapsa japonica 258 – Heterosigma akashiwo 258 Capillary electrophoresis (CE) 59 Carbon dioxide 26 Carnivorous fish 289 Cell lysis 327 Certified reference standard 458 ChemiLuminescence ImmunoAssay (CLIA) 51 Chemiluminescence multichannel immunosensor (CL-MADAG) 51 Chlorophyte 215 Ciguatera 183 – occurrence 455 Ciguatera fish poisoning 273, 428 Ciguatoxin 256, 273, 427 – action level 443 – ciguatera fish poisoning 256 – Gambierdiscus sp. 257 – Gambierdiscus toxicus 256 – Siganus rivulatus 257 Climate 181 Climate change 182, 245, 246 – environmental factor 246 – extreme events 245 – ocean acidification 243

484 | Index

– pH 246 – salinity 246 – seafood safety 245 – temperature 246 – water level 246 Climate change impacts on Alexandrium 141 Climate-ocean interaction 425 Climatological monthly average 5 CO2 138 Cochlodinium polykrikoides 185 Coherent sexual system 380 Competition 206 Continuous Plankton Recorder 40 Control and management 313 Coral 426 – Mediterranean coral 426 Critical operation point 339 Crustacean 428 Cyanobacteria 196, 203 – biomass 203 – bloom 203 – buoyant 207 – diversity 211 – filamentous cyanobacteria 426 – freshwater cyanobacteria 426 – growth 208 – Microcystis aeruginosa 205 – strain 205 Cyanobacterial bloom 313, 325 Cyanobacterial management 320 Cyanobacterial toxin 314 – cyanotoxin 314 Cyanotoxin 156 Cyclic imine 253 – fast acting toxicity 253 – gymnodimine (GYM) 253 – pinnatoxin (PnTX) 253 – pteriatoxin (PtTX) 253 – spirolide (SPX) 253 Cyclic imines poisoning 428 Cyclic polyether ladder 284 Cylindrospermopsin 156, 157, 166, 427 Cylindrospermopsis 161, 206 Cylindrospermopsis raciborskii 215 Cyst formation 133 Degradation 328 Destratification 196, 199

Detection technique – limitations 458 Diarrheic shellfish poisoning 428 Diarrheic shellfish toxin (DSP) – occurrence 456 Diatom 215, 423 – chain-forming diatom 426 Differential pulse voltammetry (DPV) 56 Dinoflagellate 287, 423 – benthic dinoflagellate 428 – dinoflagellate toxicity 427 – HAB-forming dinoflagellate 424 Dinophysistoxin 430 Directive 91/492/EEC 248 – Amnesic Shellfish Poisoning 248 – azaspiracid 248 – Diarrheic Shellfish Poisoning 248 – dinophysistoxin 248 – lipophilic toxin 248 – okadaic acid 248 – Paralytic Shellfish Poisoning 248 – pectenotoxin 248 – yessotoxin 248 Disease 417 – disease emergence 419 – foodborne disease 417 – potential disease 421 – waterborne disease 417 – Zoonotic disease 418 Dissolved organic nutrient 133 Dissolved oxygen 28 Diversity 211 Dolichospermum 161 Domoic acid 427 Drinking water 448, 449 Drug 418 – veterinary drug 419 Earth Observation (EO) 36 Ecology 325, 338 Ecosystem service 221 El Niño/Southern Oscillation 425 Electrochemical impedance 56 ELISA 302 Emerging hazard 435 Emerging species – ciguatoxin 247 Emerging toxin 244, 247 – HAB 244

Index |

– invasive marine species 244 – pinnatoxin 247 ENSO 20 Environment 426 – acidic environment 426 – environmental impact 428 – marine environment 428 Environmental factor 203 – CO2 208 – light 204 – nitrogen 201–205, 207–209, 216 – phosphorus 203–205, 207–209, 216 – temperature 196, 204 Environmental Sample Processor 38 Environmental sex determination (ESD) 381 Enzyme-Linked ImmunoSorbent Assay (ELISA) 39 Epidemiologic study 433 Estuaries 209 EU regulatory level 431 EU-Reference Laboratory 45 European Food Safety Authority (EFSA) 248 – brevetoxin 248 – ciguatoxin 248 – cyclic imine 248 – maitotoxin 248 – palytoxin 248 – tetrodotoxin 248 Eutrophic 325 Eutrophication 213 Fish 289, 418 – carnivorous fish 429 – consumable 429 – contaminated fish 421 – fish farmer 423 – fish protein 422 – fish species 421, 423 – fish stock 421 – herbivorous fish 429 – poisonous fish 432 Fishery 419 – coastal fishery 422 – commercial fishery 428 – fisheries governance 422 – global fishery 421 – local fishery 421 – small-scale fishery 422 Flow cytometry 336

485

Fluorescein diacetate (FDA) 337 Fluorescence in situ hybridization (FISH) 65 Fluorescence resonance energy transfer (FRET) 60 Food 417 – animal food 423 – aquatic food 417 – aquatic food safety 428 – contaminated food 418 – food chain 417 – food poisoning 429 – food preparation 424 – food production 220, 417 – food quality 417 – food safety 417 – food safety measure 435 – food security 420 – food web 419 – healthy food 419 Fossil cyst 468 Freshwater 423 – freshwater ecosystem 427 Freshwater food 455 Gambierdiscus 273 Gastropod 470 Genital ridge 380 Genotypic sex determination (GSD) 381 GEOSAT 15 Germinal stem cell (GSC) 381 Germline cell 380 Glacier 16 Global Alliance of Continuous Plankton Recorder Surveys (GACS) 41 Global climate change 359 – anthropogenic CO2 emission 368 – average air temperature 365 – calcifier organism 375 – carbon dioxide 368 – changes in phytoplankton communities 363 – cooperative interactions 399 – critical time element 400 – emission scenario 369 – global warming 366 – global water cycle 373 – greenhouse gas (GHG) 367 – natural carbon cycle 375 – ocean acidification 363 – ocean net primary production (NPP) 378

486 | Index

– oceans sea level 365 – oxygen concentration in seawater 376 – Oxygen Minimum Zone (OMZ) 375 – primary productivity 364 – radiative forcing (RF) 369 – reference organism 402 – Representative Concentration Pathway (RCP) 369 – temperature rise 363 Global warming 1, 196 Global water crisis 313 Globular species 274 Gonad 383 – estrogen-dependent system 382 – genetic system DMRT1 382 – genetic system SRY 382 – germline component 383 – gonad somatic sex 383 – somatic component 383 Gravity Recovery and Climate Experiment 16 Guanine-assembled graphene nanoribbons (GGNRS) 57 GYM 253 – Karenia selliformis 253 Gymnodimine 427 H2 O2 329 HAB 239 – fish kills 239 – HAB species toxicity 427 – human health 239 – prediction of HABs 433 – warm water HAB species 425 Halocline 19 Harmful algal bloom (HAB) 26, 360, 421 – amnesic shellfish poisoning (ASP) 361 – azaspiracid shellfish poisoning (AZP) 361 – ciguatera fish poisoning (CFP) 360 – cyanobacterial harmful algal bloom 421 – CyanoHAB 364 – diarrheic shellfish poisoning (DSP) 361 – neurotoxic shellfish poisoning (NSP) 361 – paralytic shellfish poisoning (PSP) 361 Health 424 – faunal health 432 – human health 428, 429 – public health costs 428 Heat wave 195 Herbivorous fish 289

High performance lateral flow immunoassay (HP-LFIA) 50 High Performance Liquid Chromatography (HPLC) 44 Homo-anatoxin-a 157 Hordeum vulgare L. 335 Human activity 242 – global impact 242 – Lessepsian migration 242 Human erythrocyte lysis assay 288 Hydraulic performance 339 Hydrodynamics 338 Hydrogen peroxide (H2 O2 ) 326 Ice area 5 Ice cover 1 – ice cover minimum 1 – pan-Arctic ice cover 8 Ice growth 9 Ice sheet 2 Ice-albedo feedback 1 Ice-free ocean 3 Imaging FlowCytobot 40 Incoherent sexual system 380 INESC-MN 53 Inflow 196 Infrastructure 219 Inhibition of enzyme activity 337 Inland water body 196 Inter-annual change 7 International Iberian Nanotechnology Laboratory-INL 53 International Union of Pure and Applied Chemistry (IUPAC) 47 Karenia 273 Karlotoxin 427 Label-free 75 Lake 196 – temperature 204 Lake Taihu 336 Lateral Flow Immunoassay (LFIA) 50 LC-MS 289 Legislation implementation 458 Lessepsian migration 249, 467 – L. lagocephalus 250 – L. suezensis 250 – Lagocephalus sceleratus 250

Index |

Lipopolysaccharides (LPS) 158 Liposoluble 284 Liquid Chromatography coupled to tandem Mass Spectrometry (LC-MS/MS) 44 Localized surface plasmon resonance (LSPR) 54 Long-term plankton record 188 Longitudinal Plasmon Band (LPB) 55 Loop-mediated isothermal amplification (LAMP) 65 Lowest observable adverse effect level (LOAEL) 441, 442 – azaspiracid 442 – berevetoxin 442 – ciguatoxin 442 – domoic acid 442 – okadaic acid 442 – palytoxin 442 – pectenotoxin 442 – saxitoxin 442 – yessotoxin 442 Maghemite 55 Magnetite Fe3 O4 55 Magnetoresistive biosensor 53 Maitotoxin 273, 427 Management framework 321 Management plan 466 Marine biotoxin 361 – anthropogenic eutrophication 362 – ballast water 362 – cultural eutrophication 362 – El Niño Southern Oscillation (ENSO) 362 Marine environment 2 Maximum Permitted Level 429 Maximum toxin level 443 mcy gene 204 Mediterranean Sea 241 – eutrophication 241 – harmful biotoxin 241 Micro-total analysis system (μ-TAS) 48 Microalga 424 – microalgal monitoring program 433 Microcystin 156, 164, 204, 427 – biodegradation 219 – gene expression 218 – production 205 Microcystis 160 Microcystis aeruginosa 205, 335 Mitigation 315

487

Mixing see Stratification, 215, 328 Molecular biology 35 Mollusk 428 – bivalve mollusk 432 Monitoring program 466 Morphology 126 Mouse bioassay (MBA) 44, 429 Mussel 425, 431 Mytilid 430 NANOMAG 77 Nanomaterial 36 – carbon nanotube (CNT) 48 – dendrimer 48 – gold nanoparticle (AuNP) 48 – magnetic nanoparticle (MNP) 48 – molecularly imprinted polymers (MIPS) 48 – Organic Framework (OF) 76 – quantum dot (QD) 48 – Single-walled carbon nanohorn (SWNH) 64 – Space-resolved SPME 76 Nanotechnology 35 Near-UV 54 Neurotoxic shellfish poisoning 428 Neurotoxin 428 New/emerging harmful species 247 – Gambierdiscus 247 – Ostreopsis ovata 247 – Vulcanodinium 247 Next-generation sequencing (NGS) 66 Noctiluca scintillans 185 Nodularia 206 Nodularin 156, 166, 427 Non-analytical method – cross-reactivity 465 – limitations 458 Nostoc 206 Nuclear magnetic resonance (NMR) 52, 302 Nucleic acid sequence-based amplification (NASBA) 70 Nutrient 199 – ammonium-nitrogen (NH+4 –N) 201 – iron 205 – N : P ratio 207 – nitrate-nitrogen (NO−3 –N) 201 – nitrogen 201–205, 207–209, 216 – phosphorus 203–205, 207–209, 216 – TN : TP ratio 204

488 | Index

Ocean 426 Ocean acidification 23, 187, 426 Ocean color 26 Ocean’s salinity 28 Oceanographic-epidemiologic study 435 Official detection method 446 Official method 429 Okadaic acid 425 Oogenesis 381 – oogonia 381 Ostreocin 468 Ozone hole 10 Pacific Decadal Oscillation 425 Palytoxin 251, 427 – benthic dinoflagellates 251 – O. mascarenensis 251 – O. ovata 251 – occurrence 457 – Ostreopsis 251 – Ostreopsis siamensis 251 – Ostreopsis spp. 251 – ovatoxin 251 Palytoxin poisoning 428 Panel on Climate Change (IPCC) 41 Paralytic shellfish poisoning 428 Pathogen 418 – foodborne pathogen 418 Pectenotoxin poisoning 428 Perennial ice 1 Performance 340 pH 26, 199 Phormidium 163 Photosynthetic pigment 37 – carotenoid 37 – chlorophyll 37 – phycobiliprotein 37 Phycotoxin 421, 424 Phylogenetic 129 Physicochemical characteristic 2 Phytoplankton 205, 421 – phytoplankton biomass 427 – phytoplankton production 422 – phytoplanktonic bloom 421 Phytoplankton bloom 240 – anthropogenic-induced stress 240 – ecological health 240 Piezoelectric-excited millimeter-sized cantilever (PEMC) 52

Plankton 20 Planktothrix 162 PnTX 253 – ester metabolite 254 – Vulcanodinium rugosum 254 Polar environment 2 Polar organic chemical integrative sampler (POCIS) 42 Polar waters 137 Polyketide synthase 287 Precipitation 216 Predictive modeling 434 Prevention 315 Prevention strategy 316 Primordial germ cell (PGC) 380 – germ cell determination 381 – germ cell sex-differentiation 381 – germ line 381 – PGC sexual differentiation 381 Propidium iodide (PI) 337 Prorocentrolide 254 – Prorocentrum lima 254 Pufferfish 432 Pyrodinium bahamense 182 Quantification cycle (Cq) 73 Quantitative Polymerase Chain Reaction (qPCR) 39 Quantitative reverse transcription PCR (qRT-PCR) 70 Quartz Crystal Microbalance (QCM) 51 Rainfall –effect on 199 – conductivity 202 – DOC 200 – lake temperature 199 – nutrients 200 – pH 202 – salinity 202 – stratification 199 – turbidity 199 – water level 200 – water residence time 200 Rainfall pattern 196 – dry period 201 – frequency 198 – intensity 200 – total amount of rainfall 198 Razor clam 428

Index |

Record low 1 Recreation 221 Recreational water 448, 452 Regulatory limit 430 Reproductive cycle 378 – gametogenesis 393 – gonad activation 393 – gonadal recession 393 – photoperiod 396 – regulation by endogenous factors 396 – regulation by exogenous factors 394 – reproductive strategy 393 – reproductive system 393 – spawning 393 – temperature 396 Reproductive period 378 Reservoir 209 Residence time 339 Ribosomal DNA (rDNA) 66 Ribosomal RNA (rRNA) 71 Ribulose-1,5-biphosphate carboxylase/oxygenase (RuBisCO) 66 Risk assessment 320, 422, 443 Risk management 433, 466 Salinity 139, 199 Sandwich Hybridization Assay (SHA) 39 Satellite data 7 Saxitoxin 129, 156, 157, 167, 427 Scallop 428 Sea 423 – Mediterranean Sea 431 – North Sea 424 Sea level 1 – global mean sea level 12 Sea surface temperature (SST) 37 Seafood 423 – seafood contamination 431 SELEX (systematic evolution of ligands by exponential enrichment) 75 Sensor 3 – satellite passive microwave sensor 3 Sex 379 – current functional sex 390 – genetic sex 379 – germline sex 379 – primary sex determination 380 – sex change 384, 385 – sex determination 380

489

– sex differentiation 380 – sex reversal 384 – sexual identity 381 – sexual phenotype 379 – somatic sex 379, 380 Sexual phase 385 – functional sexual phase 389 Shellfish 359, 419 – Aequipecten opercularis 385 – bivalve mollusk 359 – Chlamys varia 385 – Crassostrea gigas 385 – filter-feeding shellfish 424 – Mitylus galloprovincialis 385 – Ostrea edulis 385 – Pecten maximus 385 – shellfish flesh 432 – shellfish meat 432 Ship ballast water transport 184 Sir Alister Hardy Foundation for Ocean Science 40 Sludge accumulation 338 Small subunit (SSU) 66 Social sector 219 Sodium channel activator 283 Sodium channel blocker 469 Solid Phase Micro-Extraction (SPME) 43 Solid-Phase Adsorption Toxin Tracking (SPATT) 42 Spermatogenesis 381 – spermatogonia 381 Spin valve (SV) 53 Spirolide 427 SPX 253 – A. peruvianum 253 – Alexandrium ostenfeldii 253 Square Wave Stripping Voltammetry (SWSV) 62 Standard operating procedure 446 Stoichiometry 206 Stratification 196, 199, 328 Stronger wind 10 Subtropicalization 468 Summer melt 7 Surface Plasmon Resonance 49 Surface temperature 19 sxtA 132 sxtG 132 Symptom 428

490 | Index

Taxonomy 126 Temperature 196, see Environmental factor, 204, 419 – air temperature 204 – average temperature 424 – external temperature 419 – internal temperature 419 – sea surface temperature 429 – water temperature 204 Tetrodotoxin 249 – Alteromonas 249 – Charonia lampas lampas 249 – endo-symbiotic bacteria 249 – “fugu” consumption 251 – Fugu niphobles 250 – Lagocephalus lunaris 250 – Mytilus galloprovincialis 249 – occurrence 456 – producer 469 – Pseudomonas sp. 249 – Takifugu rubripes 250 – Vibrio sp. 249 Thermal adaptation 135 Thermal decomposition 329 Thermal expansion 14 Tide gauge 12 Tolerable daily intake (TDI) 441, 447 – MC-LR 447 TOPEX 15 Total Suspended Solid 28 Tourism 221 Toxicity equivalency factor 447, 462 – azaspiracid 462 – brevetoxin 464 – ciguatoxin 463 – diarrheic shellfish poisoning (PSP) 462 – microcystin 464 – oral toxicity 465 – palytoxin 464 – paralytic shellfish poisoning (PSP) 462 – tetrodotoxin 463 Toxin see Microcystin, 313, 421 – cyanobacterial toxin 421

– microalgal toxin 429 – natural toxin 423, 428 Toxin equivalent 461 Toxin production 203 Treatment method 328 Turbidity 199 Ultra Performance Liquid Chromatography (UPLC) 44 US Environmental Protection Agency (EPA) 47 US Food and Drug Administration (FDA) 46 Validation 461 Vector 424 WASH 220 Waste stabilization pond 323, 338 Wastewater stabilization pond 313 Water 417 – contaminated water 418 – deep water 422 – drinking water 421 – estuarine water 431 – inland water 422 – nutrient-rich water 425 – surface water 421 – tap water 421 – water column 425 – water extraction 418 – water flow 423 – water quality 419 – water renewal 425 – water source 418 – water species 421 – water temperature 418 Water resource sector 219 Water-soluble 287 Wet deposition 199 Wind 200, 425 World Health Organisation (WHO) 58 Yessotoxin shellfish poisoning 428 Zooplankton 425