Oceans and Health:: Pathogens in the Marine Environment

Oceans and Health: Pathogens in the Marine Environment

Oceans and Health: Pathogens in the Marine Environment Edited by

Shimshon Belkin Hebrew University of Jerusalem Jerusalem, Israel

and

Rita R. Colwell University of Maryland College Park, Maryland

Library of Congress Cataloging-in-Publication Data Oceans and health: pathogens in the marine environment/[edited by] Shimshon Belkin, Rita R. Colwell. p. cm. Includes bibliographical references and index. ISBN 0-387-23708-9 (hbk: alk. paper)—ISBN 0-387-23709-7 (eBook) 1. Marine microbiology. 2. Pathogenic microorganisms. I. Belkin, Shimshon. II. Colwell, Rita R., 1934– QR106.O244 2005 579
.165
09162—dc22 2004062636

ISBN-10: 0-387-23708-9 eBook ISBN: 0-387-23709-7 ISBN-13: 978-0387-23708-4  C 2005 Springer Science+Business Media, Inc. All rights reserved. This work may not be translated or copied in whole or in part without the written permission of the publisher (Springer Science+Business Media, Inc., 233 Spring Street, New York, NY 10013, USA), except for brief excerpts in connection with reviews or scholarly analysis. Use in connection with any form of information storage and retrieval, electronic adaptation, computer software, or by similar or dissimilar methodology now known or hereafter developed is forbidden. The use in this publication of trade names, trademarks, service marks and similar terms, even if they are not identified as such, is not to be taken as an expression of opinion as to whether or not they are subject to proprietary rights.

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(BS/DH)

Shimshon Belkin would like to dedicate this book to Miki, Yael, and Avner Belkin Rita R. Colwell would like to dedicate this book to Harold Colwell and John Liston

Contributors F. Xavier Abad Enteric Virus Laboratory Department of Microbiology University of Barcelona, Barcelona Spain

M. P. Caprais Ifremer, Centre de Brest BP 70, 29280 Plouzan´e France

Brian Austin School of Life Sciences John Muir Building Heriot-Watt University Riccarton, Edinburgh EH14 4AS Scotland, UK

Nandini Roy Chowdhury Department of Medicine State University of New York at Stony Brook Stony Brook, New York 11794, USA

Yael Barash Department of Molecular Microbiology and Biotechnology Tel Aviv University Ramat Aviv Israel 69978 Shimshon Belkin Institute of Life Sciences Hebrew University of Jerusalem Jerusalem 91904, Israel

Rita R. Colwell University of Maryland College Park and Bloomberg School of Public Health Johns Hopkins University Gianfranco Donelli Department of Technologies and Health Istituto Superiore di Sanit`a Rome, Italy

Albert Bosch Enteric Virus Laboratory Department of Microbiology University of Barcelona, Barcelona Spain

Gregory Doucette National Oceanic and Atmospheric Administration National Ocean Service Center for Coastal Environmental Health and Biomolecular Research 219 Fort Johnson Road, Charleston SC 29412, USA

Pietro Canepari Dipartimento di Patologia Sezione di Microbiologia Universit`a di Verona 37134 Verona Italy

Ronald Fayer United States Department of Agriculture Agricultural Research Service Environmental Microbial Safety Laboratory Beltsville, MD 20705 USA

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viii Charles P. Gerba Department of Soil, Water and Environmental Science University of Arizona, Tucson, AZ, USA Mich`ele Gourmelon Ifremer, Centre de Brest BP 70, 29280 Plouzan´e France D. Jay Grimes Department of Coastal Sciences The University of Southern Mississippi P.O. Box 7000 Ocean Springs, MS 39566-7000 USA Fran¸coise Le Guyader Ifremer, Centre de Brest BP 70, 29280 Plouzan´e France Dominique Hervio-Heath Ifremer, Centre de Brest BP 70, 29280 Plouzan´e France Anwar Huq Center of Marine Biotechnology University of Maryland Biotechnology Institute Baltimore, MD and UMIACS University of Maryland College Park, MD USA Iddya Karunasagar Department of Fishery Microbiology College of Fisheries University of Agricultural Sciences Mangalore-575 002 India

Contributors

Indrani Karunasagar Department of Fishery Microbiology College of Fisheries University of Agricultural Sciences Mangalore-575 002 India Jan Landsberg Fish and Wildlife Research Institute Florida Fish and Wildlife Conservation Commission 100 8th Avenue Southeast St. Petersburg, FL 33701 USA Maria del Mar Lle`o Dipartimento di Patologia Sezione di Microbiologia Universit`a di Verona 37134 Verona Italy Jean Claude Le Saux Ifremer, Centre de Brest BP 70, 29280 Plouzan´e France Jeffrey M. Lotz Department of Coastal Sciences The University of Southern Mississippi P.O. Box 7000 Ocean Springs, MS 39566-7000 USA Luisa A. Marcelino Department of Civil and Environmental Engineering Massachusetts Institute of Technology 77 Massachusetts Avenue Cambridge, MA 02139, USA Colin B. Munn University of Plymouth School of Biological Sciences Drake Circus Plymouth PL 4 8AA United Kingdom

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Contributors

G. Balakrish Nair Director, Laboratory Sciences Division ICDDR,B: Centre for Health and Population Research Mohakhali, Dhaka 1212 Bangladesh, India James D. Oliver Department of Biology 9201 University City Blvd. University of North Carolina at Charlotte Charlotte, NC 28223 USA Robin M. Overstreet Department of Coastal Sciences The University of Southern Mississippi P.O. Box 7000 Ocean Springs, MS 39566-7000 USA Rosa M. Pint´o Enteric Virus Laboratory Department of Microbiology University of Barcelona, Barcelona Spain Martin F. Polz Department of Civil and Environmental Engineering Massachusetts Institute of Technology 77 Massachusetts Avenue Cambridge, MA 02139, USA

Firdausi Qadri International Centre for Diarrhoeal Disease Research Bangladesh (ICDDR,B), Dhaka Bangladesh, India Eugene Rosenberg Department of Molecular Microbiology and Biotechnology Tel Aviv University Ramat Aviv Israel 69978 Yael Rozen Technion—Israel Institute of Technology Technion City Haifa 32000, Israel Hillel Shuval Department of Environmental Health Sciences The Hadassah Academic College Jerusalem, Israel Caterina Signoretto Dipartimento di Patologia Sezione di Microbiologia Universit`a di Verona 37134 Verona Italy

Monique Pommepuy Ifremer, Centre de Brest BP 70, 29280 Plouzan´e France

Lester W. Sinton Water Microbiology Research Laboratory Christchurch Science Centre Institute of Environmental Science and Research (ESR) 27 Creyke Road Christchurch, New Zealand

Carla Pruzzo Department of Biology University of Genova Italy

Yoshifumi Takeda Jissen Women’s University Hino, Tokyo Japan

x Janelle R. Thompson Department of Civil and Environmental Engineering Massachusetts Institute of Technology 77 Massachusetts Avenue Cambridge, MA 02139, USA James M. Trout United States Department of Agriculture Agricultural Research Service

Contributors

Environmental Microbial Safety Laboratory Beltsville, MD 20705, USA Fran Van Dolah National Oceanic and Atmospheric Administration National Ocean Service Center for Coastal Environmental Health and Biomolecular Research 219 Fort Johnson Road, Charleston SC 29412, USA

Preface The importance of combating infectious diseases has received international attention, providing the opportunity for a multidisciplinary approach that combines medicine with other scientific and technological capabilities, notably information technology, nanotechnology, and biotechnology. In fact, it has been predicted that the future will bring a merging of these technologies with the cognitive and behavioral sciences—major forces that have the potential to balance the world’s inequities. The scientific community and world leaders must work together to use knowledge and its applications to improve the condition of the planet. The connection between infectious diseases and the oceans provides a paradigm for this perspective. A stark global context indisputably frames all human health issues in the twenty-first century: the world wide movement of people and goods. Throughout the past half century, international travel has skyrocketed; there are more than 500 million international arrivals per year. The greatest increase has taken place since the mid-1990s. The world has become integrated and global; consequently, the notion that it is possible to successfully eradicate a disease from the face of the planet has become simplistic. Infectious disease is a moving target and climate shifts will affect any disease that has an environmentally sensitive stage or vector. Recognizing signals from climate models and incorporating them into health measures can provide new opportunities for proactive—rather than reactive—approaches to public health. Thus, careful attention to the role of the oceans in human health can offer new avenues of research that will provide new means of predicting and preventing those diseases that are rooted in the environment. In this volume, pathogens in the sea are reviewed by Colin Munn, who provides a broader perspective for the topic of pathogenic microorganisms associated with the world oceans. The wide diversity of microbial pathogens, their ecology, and the methodologies for their detection are described by Janelle Thompson and her colleagues, while Albert Bosch and colleagues devote their chapter to a review of human viruses, a notoriously difficult group of microorganisms to track in the marine environment. The ubiquitous vibrios are discussed by several authors and from many aspects. The topic is reviewed in detail by Carla Pruzzo and colleagues, while in-depth discussions of Vibrio vulnificus and Vibrio parahaemolyticus are provided by James Oliver and by Balakrish Nair and colleagues, respectively. The latter chapter also explores associations with higher organisms as well as seafood safety issues. A global perspective for members of this group, including climate, seasonality, sea surface temperatures, and plankton population cycles, is addressed by Rita Colwell. Understanding causation of global cholera epidemics requires a panoramic view of the ecology of vibrios in the sea, as discussed in Chapter 12. Karunasagar and Karunasagar focus on the importance of the viable but nonculturable (dormant) stage of existence to pathogenicity. Interestingly, VBNC also plays a role in geographic distribution.

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Another group of highly significant bacterial human pathogens, often not considered in the marine context, are the Gram-positive bacteria; this group is reviewed in detail by Pietro Canepari and co-workers. Regardless of the nature and the source of the pathogens, their eventual impact on their target population will depend, to a large extent, upon a combination of chemical, physical, and biotic factors. The survival of bacterial and viral pathogens is addressed by Lester Sinton and by Charles Gerba, respectively, and at the molecular level by Yael Rozen and Shimshon Belkin. The epidemiological and engineering aspects of these factors are described by Monique Pommepuy, and the economic consequences of subsequent human infections are analyzed by Hillel Shuval. Bacteria and viruses are not the only pathogens in the marine environment. The oceans also harbor a wide variety and diversity of protists, which are only beginning to be understood as a significant component of marine ecosystems. Ronald Fayer and James Trout provide a very useful introduction to the zoonotic protists that enter this environment in feces from domesticated animals, wildlife, and/or humans. Algal blooms, another newly emerging source of harm to humans and animals, are discussed by Jan Landsberg and colleagues. While this book was originally intended to discuss human pathogens only, we have come to realize that a more complete coverage of the topic should include pathogens of other organisms as well. Three groups of marine organisms of significance to marine ecology and of human economic interest were selected: fish, corals, and aquacultured crustaceans. The diseases of these three groups are respectively reviewed by Brian Austin, Eugene Rosenberg and Yael Barash, and by Jeffery Lotz and colleagues. These three concluding chapters serve to reemphasize the many levels at which the well-being of the natural environment is intertwined with human health and human interests. The distinguished authors of the chapters of this volume bring their insight and experience to bear on the issue of the health of the world oceans, but more importantly to the reliance of human health on understanding the potential of pathogens, both autochthonous and pollutionintroduced, in the marine environment. The blue planet we inhabit as human custodians of its environment, has three-fourths of its surface area covered by oceans. If we do not understand the microbial ecology of this vast expanse, especially with regard to pathogens, both identified and as yet unrecognized, we will play a deadly game of public health roulette. We thank our co-authors for the generosity of their time and knowledge and especially for their patience. All have been rewarded, I hope, by the usefulness and timeliness of this book. Shimshon Belkin Professor of Environmental Sciences The Hebrew University of Jerusalem Rita R. Colwell Distinguished University Professor University of Maryland and Johns Hopkins University

Contents 1. Pathogens in the Sea: An Overview ........................................................ Colin B. Munn 2. Diversity, Sources, and Detection of Human Bacterial Pathogens in the Marine Environment .................................................................. Janelle R. Thompson, Luisa A. Marcelino, and Martin F. Polz

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3. Biotic and Abiotic Effects..................................................................... Lester W. Sinton

69

4. Survival of Enteric Bacteria in Seawater: Molecular Aspects ...................... Yael Rozen and Shimshon Belkin

93

5. Human Pathogenic Viruses in the Marine Environment........................................................................... 109 Albert Bosch, F. Xavier Abad, and Rosa M. Pint´o 6. Survival of Viruses in the Marine Environment ........................................ Charles P. Gerba

133

7. Zoonotic Protists in the Marine Environment........................................... 143 Ronald Fayer and James M. Trout 8. Marine and Estuarine Harmful Algal Blooms: Impacts on Human and Animal Health ............................................................................. Jan Landsberg, Fran Van Dolah, and Gregory Doucette

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9. Pathogenic Vibrio Species in the Marine and Estuarine Environment............ 217 Carla Pruzzo, Anwar Huq, Rita R. Colwell, and Gianfranco Donelli 10. Vibrio vulnificus ................................................................................. James D. Oliver 11. Vibrio parahaemolyticus—Seafood Safety and Associations with Higher Organisms ....................................................................... Firdausi Qadri, Nandini Roy Chowdhury, Yoshifumi Takeda, and G. Balakrish Nair

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12. Global Microbial Ecology of Vibrio cholerae ............................................ Rita R. Colwell

297

13. Gram-Positive Bacteria in the Marine Environment .................................. Maria del Mar Lle`o, Caterina Signoretto, and Pietro Canepari

307

14. Fecal Contamination in Coastal Areas: An Engineering Approach ............... M. Pommepuy, D. Hervio-Heath, M. P. Caprais, M. Gourmelon, and F. Le Guyader

331

15. Retention of Pathogenicity in Viable Nonculturable Pathogens .................... Iddya Karunasagar and Indrani Karunasagar

361

16. Thalassogenic Infectious Diseases Caused by Wastewater Pollution of the Marine Environment .................................................................. Hillel Shuval

373

17. Bacterial Pathogens of Marine Fish........................................................ 391 Brian Austin 18. Microbial Diseases of Corals................................................................. 415 E. Rosenberg and Y. Barash 19. Aquaculture and Animal Pathogens in the Marine Environment with Emphasis on Marine Shrimp Viruses............................................... 431 Jeffrey M. Lotz, Robin M. Overstreet, and D. Jay Grimes

Index ....................................................................................................

453

1 Pathogens in the Sea: An Overview Colin B. Munn 1.1. INTRODUCTION This book considers the importance of pathogenic microbes in the sea. Pathogens are biological agents that cause disease. Humans have lived in close association with the sea for millennia and many human communities developed in coastal regions because of their importance as a source of food from fishing and as ports for transport and trade. Today, many of the world’s largest towns and cities are on the coast and vast quantities of human waste are pumped daily into coastal waters. In the last century marine aquaculture of fish and shellfish has developed as a major source of food alongside traditional harvesting methods. As well as the traditional contact with the sea as a source of food, materials and transport, the last few decades have seen increased use of the sea for leisure, sports, and recreation. Many people spend their weekends and holidays swimming, surfing, sailing, and diving. This, combined with better diagnostic and reporting systems and media interest in environmental issues, has increased public awareness of diseases associated with the sea. Given the intended readership, it is understandable that this book concentrates on pathogens of humans. As discussed below, some of these pathogens are microbes that are normally circulating in the human population, which are introduced into the sea (most commonly via sewage) and subsequently infect other humans via accidental contact or ingestion of contaminated fish or shellfish. Key questions to be addressed in this book include the survival, spread, and ultimate fate of such introduced organisms. On the other hand, many pathogens of humans are indigenous or autochthonous marine organisms, which cause disease in humans either by infection or by the effect of toxins, but humans are clearly not the normal host of these pathogens or “target” for their toxins. Key questions to be addressed include the natural ecology of these organisms and the factors that determine their ability to produce disease in humans. This understandable focus on marine-associated human pathogens may also distract us from the important realization that all types of marine organisms are the natural hosts for diverse pathogens. In the human context, disease is defined rather easily as an “impairment of health or a condition of abnormal functioning” (Biology Online Dictionary, 2004). In humans and other higher organisms this is usually manifested by visible physiological or microscopic signs of damage to tissues. However, we know very little about the many pathogens and abnormal processes that occur in the myriad of species of organisms found in natural marine Colin B. Munn Kingdom.

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University of Plymouth, School of Biological Sciences, Drake Circus, Plymouth PL4 8AA, United

Oceans and Health: Pathogens in the Marine Environment. Edited by Belkin and Colwell, Springer, New York, 2005.

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C. B. Munn

Table 1.1. Examples of viruses infecting marine organisms. Virus family

Shape

Size (nm)

Host

Adenoviridae Baculoviridae Corticoviridae, Tectiviridae Iridoviridae Lipothrixviridae Myoviridae Phycodnaviridae Podoviridae, Siphoviridae SSV1 group

Double-stranded DNA viruses Icosahedral with spikes Rods, some with tails Icosahedral with spikes Icosahedral Thick rod with lipid coat Polygonal with contractile tail Icosahedral Icosahedral with noncontractile tail Lemon shaped with spikes

60–90 100–400 60–75 125–300 400 80–200 130–200 60 60 ×100

Fungi Crustacea Bacteria Fish Archaea Bacteria Algae Bacteria Archaea

Birnaviridae Cystoviridae Reoviridae Totiviridae

Double-stranded RNA viruses Icosahedral Icosahedral with lipid coat Icosahedral with spikes Icosahedral

60 60–75 50–80 30–45

Molluscs, fish Bacteria Crustacea, fish Protozoa

Microviridae Parvoviridae

Single-stranded DNA viruses Icosahedral with spikes Icosahedral

23–30 20

Bacteria Crustacea

Caliciviridae Leviviridae Orthomyxoviridae Rhabdoviridae

Single-stranded RNA viruses Spherical 35–40 Icosahedral 24 Various, mainly filamentous 20–120 Bullet-shaped with projections 100–430

Fish, marine mammals Bacteria Marine mammals Fish

ecosystems. In these circumstances, it is appropriate to adopt a broader definition of disease, which I will define as “a condition in organisms as a result of infection by a biological agent, or the effects of its toxic products, that threatens or impairs the normal functioning of the organism in question, or threatens its life.” This allows us, for example, to consider viruses as pathogens of bacterial or algal populations, both of which are of key importance in life processes in the sea (see Table 1.1). It is also essential to emphasize that the vast majority of marine microbes are not pathogens, either in humans or in other organisms, but play essential roles in natural processes, such as geochemical cycling or beneficial symbioses. For all of the above reasons, it is important that any consideration of pathogens in the seas should take account of this broader picture. Consequently, that is the main aim of this introductory chapter. I will present a general overview of the importance of microbes in the sea, as well as an introduction to the microbial pathogens. The emphasis in this chapter will be on pathogens of marine organisms that have profound ecological significance, but that are missing from the largely human-focused approach in this book. This chapter also attempts to identify some common themes in the sources and transmission of pathogens, the mechanisms of disease production and trends in the evolution of marine-associated disease, as well as highlight gaps in our knowledge.

1.2. MARINE MICROBES AND THEIR HABITATS The oceans cover 3.6 × 108 km2 (71% of the Earth’s surface) and contain 1.4 × 1021 l of water (97% of the total on Earth). Thus, they represent the largest ecosystem on the planet,

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Pathogens in the Sea: An Overview

Figure 1.1. Schematic illustration of ecological zones of the oceans, showing marine microbial habitats (not to scale). Reproduced from Munn (2004), with permission, Garland Science/BIOS Scientific Publishers.

accounting for over 90% of the biosphere. The sea provides a huge variety of types of habitats for microbes: (a) freely suspended in the water; (b) associated with particles; (c) in sediments; (d) on the surfaces of other organisms or inanimate objects; or (e) within the bodies and cells of other organisms (Fig. 1.1). Each of these habitat types provides a range of nutritional and environmental variables, which is reflected in enormous diversity and relative abundance of different microbial types. Both physical factors (especially temperature and solar radiation) and chemical factors (especially the concentrations of oxygen and nutrients) affect the distribution and activity of different microbial types.

1.2.1. Size and Activity of Planktonic Microbes Planktonic organisms are defined as those that occur suspended in the water column and lack sufficient powers of locomotion to resist large-scale water currents. Planktonic organisms may be classified by size (Table 1.2) and include viruses, bacteria (in the nonitalicized form, this is used here as a general term to include members of the prokaryotic domains Bacteria and Table 1.2. Classification of plankton by size. Size category Femtoplankton Picoplankton Nanoplankton Microplankton a

Size range (µm) 0.01–0.2 0.2–2 2–20 20–200

Microbial groups Viruses Bacteriaa , Archaea, some flagellates Flagellates, diatoms, dinoflagellates Ciliates, diatoms, dinoflagellates, other algae

Some filamentous Cyanobacteria and sulfide oxidizing bacteria occur in larger size classes.

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Archaea), fungi, and protists (simple eukaryotic organisms traditionally classified as protozoa and algae). Although there are some notable exceptions, such as Thiomargarita (Schulz & Jorgensen, 2001), most open ocean prokaryotes are exceptionally small, even by the usual standards of microbiology. Most are less than 0.6 µm in their largest dimension and many are less than 0.3 µm, with cell volumes as low as 0.03 µm3 (DeLong, 1999). Such extremely small cells could result from a genetically fixed phenotype maintained throughout the cell cycle or because of physiological changes associated with starvation (DeLong, 1999). When subjected to starvation, many cultured bacteria (including human and animal pathogens) undergo a genetically programmed series of events resulting in changes to the cell envelope, metabolic changes, and reduction in cell size (Kjelleberg et al., 1987; Moyer & Morita, 1989; Paludan-Miller et al., 1996; Wai et al., 1999). It has been thought that the majority of these ultramicrobacteria in the open sea were derived from larger cells, normally associated with particles of organic matter, which had undergone reductive cell division as a result of starvation (Morita, 1997). However, ultramicrobacteria express higher metabolic activity per unit of volume than such large cells, and therefore represent the main component of the oceans in terms of biomass and activity (Schut et al., 1997). Their small size is an evolutionary (genotypic) adaptation to ensure efficient utilization of scarce nutrients in their oligotrophic environment as evidenced by a small genome size (Schut et al., 1997).

1.2.2. Viability, Cultivability, and Diversity An understanding of the effects of nutrient concentrations and physical factors on the physiology of marine bacteria is important not just in terms of their role in ocean processes, but also in understanding the ecology of indigenous pathogens and the fate of pathogens and indicator organisms introduced into the sea. This relates particularly to the “viable but nonculturable” (VBNC) state. This important concept was originally introduced by Colwell and co-workers to explain the discrepancy between the number of Vibrio cholerae in coastal waters that can be visualized directly (e.g. by epifluorescence) and those that can be enumerated using plate counts (Colwell et al., 1985; see Chapter 12 of this volume). The VBNC state has subsequently been demonstrated in a wide range of bacteria in the marine environment (Colwell & Grimes, 2000). The true nature of the VBNC state has continued to provoke controversy (Bloomfield et al., 1998; Kell et al., 1998) but its importance in the ecology of diverse pathogens of humans, fish, invertebrates, and of indicator organisms cannot be overstated. This is therefore a recurring theme in several chapters in this book. It is essential to distinguish the VBNC state (i.e. a transition shown when introduced to the environment by bacteria that can normally be cultured in the laboratory) from the fact that most indigenous marine prokaryotes (probably over 99.9%) have never been cultured. Spectacular advances in our knowledge of the diversity of marine prokaryotes have been achieved by the application of direct genomic methods, in which nucleic acids are isolated directly from seawater, amplified by PCR, and sequenced (Giovannoni & Cary, 1993). During the 1990s, a range of studies were conducted by different investigators in diverse marine habitats throughout the world and databases have been established containing many thousands of sequences, particularly of 16S ribosomal RNA (rRNA) genes. Based on phylogenetic analysis of 16S rRNA gene sequences, a number of clusters (clades) of Bacteria have been found, with great genetic diversity within these broad groups (Giovannoni & Rappe, 2000). The most abundant clones do not correspond to cultured species. Alpha-Proteobacteria are the dominant types, in contrast

Pathogens in the Sea: An Overview

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to the situation with collections of cultured marine bacteria, where gamma-Proteobacteria are most commonly encountered (Giovannoni & Rappe, 2000). Similarly, analysis of ocean waters using primers directed against archaeal 16S rRNA genes has led to the conclusion that Archaea are very abundant in ocean waters (DeLong, 1992, 2001; Fuhrman et al., 1992; Fuhrman & Davis, 1997). About 20% of the pelagic picoplankton is dominated by a single clade of Crenarchaeota (Karner et al., 2001). Until recently, it was assumed that most of these planktonic Bacteria and Archaea were inherently “uncultivable” because of their adaptation to the oligotrophic environment. However, the use of methods based on the dilution techniques developed by Schut et al. (1993) and Button et al. (1998), combined with fluorescent in situ hybridization and microarray technology, has recently led to successful culture of representatives of the previously uncultivable SAR11 clade (Rappe et al., 2002). This alpha-proteobacterium (now named Pelagibacter ubique) is probably the most abundant and widely distributed organism in the oceans, accounting for 25–55% of all ocean plankton prokaryotes (Morris et al., 2002). Further application of these high-throughput methods and other novel approaches is leading to cultivation of other types previously regarded as uncultivable (Connon & Giovannoni, 2002; Zengler et al., 2002).

1.2.3. Association of Microbes with Particles It is tempting to think of seawater as a homogeneous fluid containing evenly distributed microbes and nutrients. However, much evidence now supports the idea of microscale heterogeneity in the distribution of nutrients around organisms and particles of organic matter (Fenchel, 2002; Long & Azam, 2001). The water column contains large numbers of “marine snow” particles ranging from 0.5 to a few micrometers in diameter. These aggregates form as a result of collision and coagulation of primary particles such as polysaccharides, cellular debris, and fecal material (Turley, 2002). Marine snow particles contain different microhabitats, which support large diverse populations (typically 108 –109 /ml) of aerobic and anaerobic heterotrophic protists, bacteria, and their viruses (Azam & Long, 2001). Recent evidence shows that, as particles sink, the activity of bacterial extracellular enzymes results in the leaching of a plume of dissolved organic material, which provides a concentrated nutrient source for suspended planktonic bacteria (Kiorboe & Jackson, 2001). These may show rapid motility and chemotactic behavior in order to remain within favorable concentrations (Mitchell et al., 1996; Kiorboe & Jackson, 2001) and the nature of this behavior seems fundamentally different to that in familiar bacteria, such as Escherichia coli (Mitchell, 2002). There are extensive interactions between members of the microbiota associated with marine snow particles, for example via the production of antibiotic compounds and extracellular enzymes (Long & Azam, 2001) or density-dependent quorum sensing mechanisms (Gram et al., 2002).

1.2.4. Phototrophs, Heterotrophs, and Chemoautotrophs The top 100–200 m of the water column constitutes the photic zone and different phototrophic organisms are found at different depths according to their physiological properties. Recent evidence shows that the diversity and distribution of phototrophs is much greater than previously thought. Whilst the traditionally designated “phytoplankton” (dinoflagellates and diatoms) occur near the surface and carry out photosynthesis based on chlorophyll a, phototrophic metabolism by prokaryotes is very varied. The cyanobacterial genera Prochlorococcus and

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Synechococcus account for about two-thirds of ocean productivity (Bryant, 2003). Prochlorococcus contains divinyl chlorophylls adapted to the harvesting of blue light, which penetrates to 150–200 m. There are distinct ecotypes, which colonize different depths, and these show major differences in their genome sequences (Dufresne et al., 2003; Rocap et al., 2003). Genomic techniques have also demonstrated the importance of previously unknown types of phototrophic metabolism in ocean processes, revealed by the widespread distribution of the protein bacteriorhodopsin in gamma-Proteobacteria (Beja et al., 2000, 2001) and the diversity of aerobic anoxygenic phototrophs (Beja et al., 2002; Kolber et al., 2000). However, the photic zone is a very tiny proportion of the overall volume of water in the oceans and most microorganisms are chemoheterotrophic, living via the metabolism of dissolved organic matter (Fig. 1.2). It has been estimated that the total numbers of Bacteria and Archaea in the water column of the oceans are 3.1 × 1028 and 1.3 × 1028 , respectively (Karner et al., 2001).

Figure 1.2. Simplified representation of food webs and microbial interactions in the sea. Arrows show the transfer of organic material by various routes. For clarity, not all trophic connections are shown and regeneration of inorganic nutrients by heterotrophic bacterial activity is omitted. Reproduced from Munn (2004), with permission, Garland Science/BIOS Scientific Publishers.

Pathogens in the Sea: An Overview

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1.2.5. Microbiology of Marine Sediments The ocean floor is covered with sediments originating from erosion of the land, volcanic material, and cosmic dust, combined with “oozes” derived from the accumulation of biogenic materials such as marine snow and the skeletons of calcareous (foraminifera and coccolithophorids) or siliceous (diatoms and radiolarians) protists. Ocean sediments have accumulated over millennia and may be thousands of meters deep, with buried organic material joining the geological cycle and eventually reemerging through tectonic uplifting. The microbiology of deep marine sediments and subsurface rocks is an area of active investigation (Reed et al., 2002). Water saturation restricts penetration of oxygen to the sediments, and microbial activity below the top few cm of sediments is anaerobic. Microbial processes such as methane production and oxidation or the oxidation and reduction of sulfur compounds are of particular importance. High nutrient levels in some regions result in such extensive microbial oxygen demand that the overlying water column becomes anoxic (e.g. the Black Sea and ocean basins off Venezuela and Mexico). Highly diverse communities of Bacteria, Archaea, and protists exist in the sediments and only a tiny fraction of those identified using molecular methods can currently be cultivated. One example of the surprising results from the application of new methods is the discovery of unique consortia between Bacteria and Archaea responsible for anaerobic methane oxidation (Boetius et al., 2000). Another is the discovery of Nanoarchaeum equitans, an extreme thermophile belonging to a completely new division of the Archaea (Huber et al., 2002) and which is an obligate parasite of another archaeon, Igniococcus. N. equitans has the smallest genome of any cellular organism yet sequenced (Waters et al., 2003). A better knowledge of community structure, activity, and interactions in marine sediments will be of great importance in understanding the ecology of pathogens. Many of the subsequent chapters highlight our lack of understanding of the factors that affect the survival of pathogens in sediments, which constitute a reservoir of infection by human viruses and bacteria (Chapters 3 and 5), fish viruses and bacteria (Chapter 17) or toxic dinoflagellate blooms (Chapter 7).

1.2.6. Biofilms and Microbial Mats Biofilms consist of a collection of microbes bound to a solid surface by their extracellular products, which trap organic and inorganic components (Sutherland, 2001). All kinds of marine surfaces, including rocks, animals, plants, and fabricated structures, may be colonized by biofilms (Decho, 2000). This has many important consequences for marine ecology, especially since it affects the settlement and development of invertebrate larvae or algal spores (Cooksey & Wigglesworth-Cooksey, 1995; Joint et al., 2000) and has economic importance in biofouling and corrosion (Callow & Callow, 2002; Hill, 2004). Bacteria and diatoms can both initiate biofilm formation due to environmental cues, especially nutrient availability (Cooksey & Wigglesworth-Cooksey, 1995) and a series of complex developmental changes occur in the transition from suspended planktonic microbes to the sessile, attached form (Watnick & Kolter, 2000). Detailed studies of biofilm formation by several bacterial pathogens have been undertaken, showing that the establishment of the biofilm involves complex genetic regulation of genes involved with motility and exopolymer production. For example, in the human pathogen V. cholerae genetic differences between serotypes explain differences in the rate of biofilm

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formation (Watnick et al., 2001), which may be relevant to the ecology of the pathogen during its phase as a colonizer of zooplankton and consequently, the epidemiology of human infection (Chapters 9 and 12 of this volume). Another human pathogen, Vibrio parahaemolyticus (Chapter 11), possesses multiple life stages according to environmental conditions and undergoes dramatic changes in cell morphology and flagellar synthesis during its transition from a planktonic form to biofilm cells (McCarter, 1999). The nature of the surface can have profound effects. For example, many marine bacteria express chitin binding and chitinase genes selectively when they encounter chitin (Baty et al., 2000). This can have important consequences during infections of crustacea, which often show lesions of the exoskeleton (McGladdery, 1999). When surfaces are colonized, ecological succession resulting from metabolic processes results in the development of micro-environments and mixed communities of bacteria and protists which can form mats several cm thick. These are particularly important in shallow and intertidal waters (Stal, 2001), where their composition and development is affected by light, temperature, flow rate, pH, redox potential, oxygenation, and the levels of inorganic and organic compounds. Phototrophs such as cyanobacteria and diatoms are major components of stratified mats and the species composition and zonation are determined by the intensity and wavelength penetration of light. Diurnal variations result in gradients of oxygen and sulfide and microbes may migrate through the mat to find optimum conditions at different times. An example of the importance of complex microbial mats in diseases is shown in black band disease (BBD) and some related destructive infections of coral reefs that are discussed in detail in Chapter 18. In BBD, microscopic examination reveals a consortium consisting of a large gliding cyanobacterium identified as Phormidium corallyticum and numerous heterotrophic bacteria, including Beggiatoa and sulfate-reducing bacteria (SRB) (Richardson, 1998). It is generally assumed that Phormidium initiates the disease (perhaps through production of toxins and necrotic enzymes) and that the anoxic and sulfide-rich conditions generated at the base of the band result in the death of the coral tissue. However, molecular analysis of the communities associated with healthy and diseased corals has produced differing results (Cooney et al., 2002; Frias-Lopez et al., 2002) and the etiology of BBD remains unclear.

1.2.7. Mutualistic Symbiotic Interactions The marine environment contains innumerable examples of symbiotic interactions between organisms. The term symbiosis (literally “living together”) actually refers to any permanent relationship between two different organisms, which can range from loose commensal associations to permanent, obligate parasitic interactions, although in common usage the term is usually synonymous with mutualism, in which both partners benefit. By brief reference to a few examples of symbiotic associations involving microorganisms at different levels of biological organization, I will illustrate the relevance of this field to the subject of pathogens in the sea. At the first level, there is growing evidence that many marine prokaryotes do not exist independently in nature, but instead cooperate to ensure the effective utilization of nutrients through syntrophy. A recently discovered example illustrates the importance of these processes for other forms of life. Boetius et al. (2000) showed that clusters of methanotrophic Archaea surrounded by a “shell” of SRB are together capable of the oxidation of methane in anaerobic sediments. This activity, which occurs in massive reef-like mats at methane seeps, generates

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sulfide that is used by a wide variety of sulfide-oxidizing bacteria, either free-living or symbiotic in invertebrates (see below). Bacterial symbionts have been observed in a wide variety of marine protists, in particular ciliates, dinoflagellates, and diatoms, with associations ranging from loose ectosymbiotic associations, through intracellular endosymbioses to fully integrated organelles. This has great importance in nutrient cycling in different environments. For example, many ciliates in anaerobic habitats are hosts for SRB (Fenchell & Finlay, 1995) or methanogenic archaea (van Hoek et al., 2000). As a group, protists show a spectrum of nutrition ranging from fully autotrophic (phototrophy) to absolute heterotrophy by particle ingestion or phagotrophy (Caron, 2000). This is one of the reasons why traditional classification of simple eukaryotes into algae (“plant-like”) and protozoa (“animal-like”) is so flawed. This becomes especially important in the dinoflagellate group, which contains many genera that produce toxins affecting humans and other organisms and may show a spectrum from fully photosynthetic, through mixotrophic to fully phagotrophic at different stages of the life cycle (Caron, 2000). Such organisms may possess sophisticated regulatory mechanisms to control the balance of photosynthesis, uptake of dissolved nutrients and feeding by ingestion of particles (Stoecker, 1999). Knowledge of these processes will have great significance for understanding the effect of environmental conditions on the life cycle and toxin production, as discussed in later chapters (Chapters 7 and 8). A case which dramatically illustrates the complex issues involved is that of Pfiesteria piscicida, a newly identified dinoflagellate reported to cause mortalities in fish (Burkholder & Glasgow, 1997, 2001) and illness in humans (Grattan et al., 2001), where controversy surrounds the stages of the life cycle and toxin production (Litaker et al., 2002; Burkholder & Glasgow, 2002). Symbiotic bacteria may also be directly responsible for the production of dinoflagellate toxins. Many studies have shown that several bacteria can produce these toxins in pure culture (Gallacher & Smith, 1999; Lee et al., 2000), but there has been disagreement about the validity of these claims (Matsumura, 2001), and attempts to resolve the origin of these toxins will depend on the development of improved assay methods (Smith et al., 2002) and genetic analysis of bacteria associated with dinoflagellates (Plumley et al., 1999; Hold et al., 2001). This important topic is explored further in Chapter 8. Many marine invertebrates contain endosymbiotic chemoautotrophic bacteria. This phenomenon came to light with the discovery of extensive animal communities at the hydrothermal vents (Van Dover, 2000). The tissues of animals such as the tubeworm Riftia pachyptila, the clam Calyptogena magnificus, and the mussel Bathymodiolus thermophilus were shown to be packed with bacteria which provide carbon compounds to the host as a result of chemolithotrophic metabolism, using the oxidation of sulfide as a source of energy (Van Dover, 2000). Another important vent animal, Alvinella pompejana, is covered with different groups of epibiotic bacteria (Campbell et al., 2001; Cary et al., 1997), which contain key enzymes of the reductive tricarboxylic acid cycle (Campbell et al., 2003) involved in chemolithotrophic metabolism. It was initially thought that such chemolithotrophic symbioses were restricted to these specialized habitats of hydrothermal vents. However, it is now known that they occur in a very wide range of animal hosts, some (but not all) related to the vent animals, in many habitats, wherever the correct balance of reducing conditions and sulfide concentration occurs. In one such association, two types of Proteobacteria coexist within the oligochaete worm Olavius algarvensis (Dubilier et al., 2001), providing a further example of syntrophy between bacterial partners. One of the bacteria is a member of the alpha-Proteobacteria and oxidizes

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sulfide, whilst the other is a sulfate-reducing member of the delta-Proteobacteria. The mutual activities of the two bacteria enable the worm to colonize new habitats without being dependent on a continual supply of sulfide. The presence of chemolithotrophic symbionts can have great effects on the behavior of their hosts, for example affecting the burrowing of clams (Dufour & Felbeck, 2003) and has clearly been a major factor in the evolution of certain groups of marine invertebrates (Distel et al., 2000). The involvement of complex microbial communities in symbioses with marine animal hosts is further illustrated by consideration of scleractinian corals. In common with a range of other marine invertebrates, corals contain dinoflagellates (Symbiodinium sp.) within their tissues, known as zooxanthellae. The dinoflagellates carry out photosynthesis and excrete a high proportion of the fixed carbon to the coral in the form of glycerol and other compounds. This contribution to coral nutrition is essential for the rapid growth of the skeleton responsible for reef building (Glynn, 1991) and its importance is demonstrated by the phenomenon of coral bleaching, in which environmental stress leads to loss of the zooxanthellae or loss of photosynthetic pigments (Brown, 1997). Coral bleaching has increased in frequency and intensity in recent years and there is extensive evidence that climate changes (especially higher than normal sea temperatures and increased solar irradiation) are responsible. As noted by Rosenberg and Ben Haim (2002), many coral biologists do not consider bleaching to be a disease, although it clearly fits most definitions of this term. In the past few years, convincing evidence has begun to accumulate that bacterial infection may be responsible, at least in some cases, for the initiation of coral bleaching and Chapter 18 reviews recent thinking about the generality of this hypothesis. Bleaching could also possibly be caused by temperature induction of latent viruses in the zooxanthellae (Wilson, personal communication) and preliminary evidence for this phenomenon in a sea anemone has been described (Wilson et al., 2001). Variations in susceptibility and recovery of corals to bleaching and the discovery of multiple genetic forms of Symbiodinium led to the formulation of the adaptive bleaching hypothesis (Buddemeier & Fautin, 1993), which sees the bleaching response as a natural response to environmental stress. Several studies, notably those of Rowan et al. (1997), have provided evidence in support of the hypothesis. In reviewing the current status of the hypothesis, Buddemeier et al. (2004) emphasize its ecological significance because of the potential for changing combinations of hosts and symbiotic dinoflagellates to create new “ecospecies” that have different fitness for various environmental conditions. They also emphasize the fact that coral is a “holobiont” that comprises the animal tissue, zooxanthellae, and diverse bacterial communities. Rohwer et al. (2001, 2002) and Bourne & Munn (2003, 2005) demonstrated a wide diversity of bacterial types and it is possible that specific bacteria may associate with particular corals and undoubtedly influence the health and nutrition of the animal host. Given the fact that some of these coral-associated bacteria have antagonistic interactions (Bourne & Munn, 2003) and that some may harbor latent bacteriophages with effects on other community members (Munn, unpublished), the complexity of these symbiotic interactions becomes even more apparent. The biology of symbiotic interactions between bacteria and sponges has also attracted considerable interest. This is mainly because sponges are a rich source of antimicrobial compounds and growing evidence suggests that these originate from bacterial symbionts, playing a role in the defense of their host sponge (Webster et al., 2001). Another important example of a mutualistic symbiotic interaction between a microbe and an animal host is the colonization of the light organ of the bobtail squid Euprymna scolopes by bioluminescent Vibrio fischeri. Once the association is fully developed, dense populations

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of V. fischeri emit light, which probably provides the squid with protection from predation due to the phenomenon of counter-illumination (Ruby, 1996). The squid eliminates about 90% of the contents of the light organ each day as a thick paste and the bacterial colonists regrow until they reach a critical density, at which point they become bioluminescent (Ruby, 1996). Regulation of the expression of the bacterial genes for luminescence is controlled by densitydependent quorum sensing (Bassler, 2002; Dunlap, 1999). The nascent light organ of newly hatched squid traps the bacteria from seawater (McFall-Ngai & Ruby, 1998), which swim into crypts of the organ (Graf et al., 1994) and influence its subsequent development (Foster et al., 2000; McFall-Ngai & Ruby, 2000). Various classes of mutants exist that are able to initiate infection, but do not grow or persist in the light organ. For example, bacteria must be able to utilize the rich mixture of amino acids secreted by the host (Graf & Ruby, 1998). Among the many interesting discoveries is the fact that mutants defective in the key bioluminescence genes luxA, luxI, and luxR fail to initiate the host cell differentiation which characterizes the symbiotic infection (Visick et al., 2000). It is difficult to explain why luminescent strains out-compete nonluminescent mutants unless luminescence is beneficial to the bacterium as well as the host. Ruby & McFall-Ngai (1999) propose that luminescence may have evolved as a mechanism for controlling oxygen levels and the avoidance of oxygen stress in the light organ. Additional evidence for this hypothesis is provided by the observation that an inducible catalase, which breaks down hydrogen peroxide, is necessary for normal colonization (Visick & Ruby, 1998). Ruby and McFall-Ngai (1999) note that V. fischeri produce proteins with ADP-ribosylating activity which may also interfere with the oxidative burst in the host tissue and postulate that similar products such as the cholera toxin (see Chapter 9 of this volume) might have evolved to promote benign relationships with unknown aquatic hosts, rather than as toxins. In conclusion, this brief review of examples of marine symbioses illustrates that the distinctions between mutualistic symbiosis and pathogenesis and between health and disease are less obvious than we might at first think. Mechanisms such as gene transfer, quorum sensing, and signaling/sensing systems emerge as common themes in many bacterial symbionts and pathogens (Hentschel et al., 2000; Steinert et al., 2000).

1.2.8. Pathogens of Marine Planktonic Microbes Bacteriophages belonging to a number of different virus groups (Table 1.1) infect bacterial and archaeal cells, playing a critical and central role in ocean food webs. They are responsible for about 10–50% of bacterial mortality (Fuhrman, 1999; Fuhrman & Noble, 1995), leading to recycling of organic matter in microbial loop processes and contributing to the microscale heterogeneity of seawater through the release of polymeric substances (see Fig. 1.2). Bacteriophages infecting cyanobacterial and other phototrophs have major effects on the dynamics of photosynthesis in the oceans and hence affect primary productivity. For example, cyanophages have recently been identified (Sullivan et al., 2003) for the different light-adapted ecotypes of Prochlorococcus, which are the dominant photosynthetic organisms in tropical and subtropical oceans, making a major contribution to global carbon dioxide fixation (Goericke & Welschmeyer, 1993; Liu et al., 1997). Proctor & Fuhrman (1990) have shown that 1–3% of marine cyanobacterial cells contain assembled viruses and are apparently nearing the end of the lytic cycle. As with heterotrophic bacteria, it is thought that the biggest impact of virus infection on cyanobacteria is to affect the species or strain composition of the community, rather than total mortality (Fuhrman, 1999).

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The importance of virus infection of eukaryotic phytoplankton has also only recently been appreciated and is an area of intense current research. The best known examples to date are members of the Phycodnaviridae, a group with some of the largest genomes known (Wilson, 2003). Infection of the coccolithophorid Emiliania huxleyi has been shown to be due to a newly identified double-stranded (ds) DNA virus (Bratbak et al., 1993; Schroeder et al., 2002), which is responsible for the sudden collapse of the large summer blooms of E. huxleyi that occur in nutrient-rich coastal waters (Brussaard et al., 1996). A number of other planktonic eukaryotic microbes have also been shown to host viral pathogens (Cottrell & Suttle, 1995; Suttle & Chan, 1995; Milligan & Cosper, 1994). Of particular significance is the role of viral lysis in the release of dimethylsulfide propionate (an osmoprotectant compound found in many types of algae). The subsequent conversion of this compound to dimethyl sulfide has many effects on ocean ecology and atmospheric processes, most importantly acting as nuclei for cloud condensation and leading to an albedo effect with subsequent effects on global climate (Hill et al., 1998). Besides its importance in general ecological processes in the sea, viral lysis of microalgae has particular importance in the fate of harmful algal blooms (HABs). As noted above, and discussed in detail in Chapter 8, proliferation of toxic species in HABs can result in disease in marine life (e.g. fish, birds, and mammals) and humans. The possibility of controlling the proliferation of toxic HABs (as well as deleterious nontoxic blooms) by viral infection is an area of current interest (Nagasaki et al., 1999). Tarutani et al. (2000) monitored the population dynamics of blooms of the raphidophyte alga Heterosigma akashiwo (which forms blooms toxic to fish) infected with a large ds DNA virus (member of the Phycodnaviridae). Total abundance and clonal composition was affected by viral infection. The discovery of HaRNAV, a single-stranded (ss) RNA virus that also infects H. akashiwo (Tai et al., 2003), emphasizes the diversity of virus pathogens of algae and the complexity of virus–host interactions in the environment. Distinct ds DNA and ss RNA viruses have also been identified as pathogens of the bivalve-killing dinoflagellate Heterocapsa circularisquama (Tarutani et al., 2001). A search for virus pathogens of other HAB-forming species could be very profitable in attempts to control their impacts.

1.2.9. Pathogens of Seaweeds and Marine Plants In view of the high numbers of microbes in the ocean, it is surprising that diseases in marine macroalgae appear to be relatively rare. There is no doubt that this can be partly explained by the fact that algal pathology is a neglected area of research (Andrews, 1976), but it seems that algae possess effective defense mechanisms against microbial colonization and infection (Kubanek et al., 2003). Harvesting of seaweeds for foods, nutraceuticals, and products such as carageenins and agar is economically important in many parts of the world. Intensive aquaculture of certain species of red, brown, and green algae now constitutes the major source of seaweed products in several countries including China, Japan, Indonesia, The Philippines, Korea, and Chile. Pythium porphyrae (a phycomycete fungus) is responsible for a wasting disease resulting in extensive coalescing lesions of the red alga Porphyra, harvested as the food crop “nori” in Japan (Andrews, 1976; Amano et al., 1996). A range of other phycomycetes has been associated with infections of various seaweeds. Ascomycetes also cause infections of Sargassum, Laminaria, and other kelps (Andrews, 1976). The red seaweed Kappachus alvarezii, which is the most important farmed species in The Philippines for the production of carageenins, is susceptible to “ice-ice disease,” a phenomenon caused by bacterial colonization of the algal fronds if stressed

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by low salinity, high temperature, or poor light intensity. Opportunistic Vibrio spp. have been identified as the causative agent (Largo et al., 1998, 1999). Infection of the red alga Mazzaella laminaroides by an unknown agent appears to exert its main effect by causing deformation, which increases susceptibility of the algal fronds to grazing (Buschmann et al., 1997). Among the limited recent reports of infections in natural populations of macroalgae are descriptions of diseases of coralline algae in the Indo-Pacific caused by unidentified colonial bacteria (coralline lethal orange disease; Littler & Littler, 1995) and fungi (Littler & Littler, 1998). As well as the direct damage caused by infection of coral animals discussed in Chapter 18, infection of coralline algae also has a great influence on the formation and ecology of reefs. Algae can themselves be pathogenic to other species of algae. For example, the green alga Ulva rigida is susceptible to infection by the endophytic filamentous alga Aerochaete geniculata, resulting in degradation leading to secondary bacterial infections (del Campo et al., 1998). Oligosaccharide signals have been shown to be important in cell–cell recognition in infection of another cultivated green seaweed, Chondrus crispus, by A. geniculata (Bouarab et al., 2001). The microscopic brown algae Laminarionema elsbetiae and Streblonema aecidioides have been shown to be endophytic pathogens of kelps (Laminaria spp.), causing severe morphological changes to the thallus (Peters & Schaffelke, 1996; Heesch & Peters, 1997). The mechanisms by which algae resist colonization by pathogens have been the subject of recent research, driven largely by the quest to discover novel pharmaceuticals and new methods to prevent biofouling. Among the most important discoveries is the finding that the red alga Delisea pulchra produces halogenated furanones, a class of compounds shown to inhibit acyl homoserine lactone- mediated quorum sensing signaling systems in bacteria (Kjelleberg et al., 1997). This affects the overall bacterial colonization and species composition of the seaweed surface. In another recent study, the novel compound lobophorolide was isolated from the brown alga Lobophora variegata and found to have potent specific activity against saprophytic and pathogenic fungi (Kubanek et al., 2003). The normal microbiota of seaweed surfaces may also protect them from colonization and infection. Pigmented marine bacteria in the genus Pseudoalteromonas (Egan et al., 2001), which populate the surface of the green alga Ulva lactuca, have been shown to produce a range of extracellular compounds that inhibit colonization by other algae, bacteria, and fungi, as well as invertebrate larvae (Egan et al., 2002). Further investigation of the natural defense mechanism of seaweeds is likely to provide more insight into the relatively low incidence of microbial disease, as well as yield potentially useful products for biotechnological exploitation. Disease also affects the productivity of higher plants found in shallow marine habitats, especially seagrasses (Zostera, Thalassia, and related species). This can have considerable ecological importance, since seagrass beds are major habitats for fish and other marine life, including endangered species such as turtles and dugongs, as well as feeding grounds for ducks and geese. Since the 1920s, there have been periodic reports of infection of large areas of seagrass beds (“wasting disease”) in Northern Europe and the USA (Roblee et al., 1991). Such epidemics have virtually eliminated Zostera in many locations. As with many protists, the taxonomic position of the causative organism has been uncertain and it was first described as a fungus or slime mold (Muehlstein et al., 1988). Phylogenetic studies using rRNA gene sequencing now indicate that Labyrinthula forms a monophyletic group within the major eukaryotic division Heterokonta (Honda et al., 1999) and the species Labyrinthula zosterae has been shown to be a primary pathogen responsible for inhibition of photosynthesis and necrosis (Ralph & Short, 2002).

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In concluding this section, it is noteworthy that no virus infections appear to have been reported in seaweeds or marine plants.

1.2.10. Diseases of Invertebrates The marine invertebrates are among the most diverse groups of animal life on Earth, with hundreds of thousands of species. Knowledge about microbial infections is only available for a tiny fraction of these, almost exclusively those of economic importance as food animals. In particular, bivalve molluscs and crustacea have been harvested from the sea and cultured in traditional extensive systems for millennia and diseases of these animals are briefly reviewed in the following sections. 1.2.10.1. Bivalve Molluscs Whilst a few natural epizootics have been described, the most severe virus infections have occurred as a result of movement of animals to new geographic areas and intensive aquaculture practices (McGladdery, 1999). In the 1960s, a major epizootic of gill necrosis caused by an iridovirus occurred in France in the oyster Crassostrea angulata and virtually destroyed the European oyster fishery. Replacement of stock by Pacific oysters (Crassostrea gigas) was followed by further infection. Birnaviruses and herpes viruses have been isolated as agents of disease in several species of bivalve molluscs (Arzul et al., 2001) although proof of etiology and study of pathogenesis is usually lacking, except in cultured species. Probably, the most important cause of disease in cultured bivalves is bacterial infection of larvae in the hatchery stage. Members of the genus Vibrio (especially V. anguillarum, V. tubiashi, and V. alginolyticus) are most commonly implicated (McGladdery, 1999) although full identification is often incomplete or inaccurate because of difficulties in interpretation of biochemical and other diagnostic tests. Growth of pathogens is encouraged by deterioration in hatchery water quality due to the accumulation of organic matter, although bivalve larvae of different species are very variable in their susceptibility to bacterial density (Perkins, 1993; Sindermann, 1989). A variety of extracellular toxins (hemolysin, protease, and a ciliostatic factor) have been shown to be responsible for necrosis of larvae (Birkbeck et al., 1987; Nottage & Birkbeck, 1987). Bacterial infections are uncommon in adult bivalve molluscs, but a notable exception is brown-ring disease in Manila clams, caused by Vibrio tapetis, which attaches to the clam tissue causing abnormal thickening and deposition of material around the inner shell (Allam et al., 2002). Vibrio harveyi has been identified as the cause of mortalities in cultured pearl oysters in north-western Australia (Pass et al., 1987). Juvenile oyster disease causes significant losses due to seasonal mortalities of hatchery-produced oysters in the eastern USA. The disease has several similarities to brown-ring disease in clams, suggesting a bacterial etiology, and a previously unknown alpha-proteobacterial species of the Roseobacter group has been identified as the pathogen responsible (Boettcher et al., 2000). A large number of rickettsia-like pathogens have been reported in at least 25 species of bivalves (McGladdery, 1999; Wu & Pan, 1999). Generally, these cause few signs of disease and mortality in adult animals is usually low, unless there are environmental stresses such as sudden temperature change. However, as with other bacteria, the larvae are very susceptible to these obligate intracellular pathogens, but little is known about pathogenicity mechanisms.

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1.2.10.2. Crustacea Viral diseases in intensively cultured crustacea (especially penaeid shrimps) have emerged to become the most important factor limiting further development of the industry. The spread of disease is favored by intensive methods in hatcheries, the use of wild-caught broodstock, and the widespread transport of eggs, larvae, mature animals, and contaminated products across geographic boundaries, as well as the cannibalistic feeding habits of many species. Although some of these agents were first recognized as agents of disease in wild crustaceans (Munday & Owens, 1998), the number of viruses recognized as important in shrimp aquaculture has risen from 6 in 1988 to nearly 20 today, with the expansion of range and volume of intensive culture. Many shrimp viruses are members of the baculovirus group (Table 1.1). Lightner and Redman (1998) emphasize the importance of advances in the development of molecular methods for the diagnosis of shrimp diseases. The most well known bacterial disease of crustacean is gaffkaemia in lobsters (Homarus americanus and Homarus gammarus), caused by Aerococcus viridans var. homari. The disease is highly contagious and gains entry to the hemolymph of the animal via abrasions in the cell and is probably acquired from a reservoir in wild animals, with disease spread linked to the transport of animals across large distances (Alderman, 1996). Intracellular rickettsias and mycoplasmas also cause epizootics with high mortalities in crustacea such as crabs, lobsters, and penaeid prawns, usually by infection of the hepatopancreas (McGladdery, 1999). Such epizootics can have marked effects on marine ecology, for example those caused by currently unidentified rickettsias infecting crabs (Cancer and Carcinus spp.) in Europe and the USA. Various Vibrio spp. cause infections of crustacea, particularly in the larval and juvenile stages. Besides their importance as the cause of devastating losses in shrimps and prawns in hatcheries and culture ponds, especially in the tropics, these diseases reveal some significant messages of relevance to the study of human pathogens. Of particular interest is the discovery of bacteriophage-mediated virulence in V. harveyi. Highly virulent strains produce two lethal exotoxins, which probably act in the larval intestine (Harris & Owens, 1999). Oakey and Owens (2002) demonstrated that a latent bacteriophage could be induced in virulent strains by mitomycin C induction. Munro et al. (2003) showed that infection of an avirulent strain with this virus resulted in increased virulence associated with the production of several new extracellular virulence products. Such bacteriophage-mediated virulence also occurs in V. cholerae (Waldor & Mekalanos, 1996; Karaolis et al., 1999) and other human pathogens and probably represents a general explanation for the movement of genes between symbionts and other bacteria in the evolution of pathogenicity (Hentschel et al., 2000). Genetic exchange may also lead to variability in biochemical identification tests for Vibrio spp. (Vandenberghe et al., 2003). Bacteriophage therapy has produced some promising results for the control of bacterial diseases in aquaculture (Flegel, 2002; Nakai & Park, 2002). The demonstration of bacteriophage mediated virulence indicates that careful evaluation is needed before this practice is widely adopted. 1.2.10.3. Corals, Sponges, and other Marine Invertebrates A variety of bacteria and fungi have now been identified as coral pathogens (Richardson, 1998) and the importance of this field continues to grow, with increasing recognition of the apparent spread in diseases correlated with pollution and global climate change (Rosenberg &

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Ben Haim, 2002). This topic is explored in detail in Chapter 18. The recent demonstration of the role of a marine invertebrate as a vector for coral disease (Sussman et al., 2003) has potentially far-reaching implications in understanding the epidemiology or epizootology of marine-associated diseases. With the exception of their possible involvement in coral bleaching noted above, the role of viruses as disease agents of corals, sponges, and other important marine invertebrates has hardly been investigated, largely because of the difficulties of study in the absence of cell cultures for propagation of suspected viral agents. These are areas ripe for further investigation. Besides corals, disease in other invertebrates in reef systems has caused major ecological disturbances. For example, in the 1990s a series of mass mortalities of various species of sponges occurred in the Florida Keys, coinciding with excessive blooms of cyanobacteria (Butler et al., 1995). One of the most dramatic mass mortalities in marine habitats occurred in 1983, when over 97% of populations of the sea urchin Diadema antillarum in parts of the Caribbean were suddenly killed, with far-reaching and long-term consequences for the ecology of coral reefs (Lessios, 1995). Although an infectious agent was suspected, this was not identified (Harvell et al., 1999). In the 1990s, sudden epizootics in the Caribbean causing mass mortalities of gorgonian sea fans were attributed to the soil fungus Aspergillus sydowii (Alker et al., 2001) as a result of the deposition of wind-blown dust from the Sahara desert (Shinn et al., 2000) and this is suspected as a cause of previous mass mortality events in the Caribbean.

1.2.11. Pathogens of Marine Fish Because of the importance of fish as a food source, fish diseases have been extensively studied. Fluctuations in the natural populations of fish due to microbial infections must undoubtedly be overlaid on the impact of over-fishing, driving some commercially important species to near extinction (Pauly et al., 2002). Fish diseases have assumed particular significance following the widespread development of intensive aquaculture since the 1970s, which has promoted the emergence of disease through stress on the host and improved opportunities for rapid transmission of the pathogen. Viral and bacterial pathogens of fish are considered in detail in Chapter 17, and only a few general points will be made here. Despite the abundance of microbial pathogens of fish, quantitative estimates of their impact on the ecology of natural populations are lacking (Essbauer & Ahne, 2001). Notable exceptions include viral hemorrhagic septicemia (VHS) virus, which is responsible for severe epizootics in populations of wild herring, pilchards, mackerel, and other species in northern Europe and Canada (Hedrick et al., 2003) and lymphocystis virus, which causes a highly contagious chronic infection resulting in hypertrophy of the skin cells (Egusa, 1992). VHS and lymphocystis, as well as a range of other virus infections including infectious pancreatic necrosis, infectious salmon anemia, infectious hematopoietic necrosis, and herpesviruses, are the cause of significant mortalities in farmed fish. A wide range of bacterial pathogens has also been described in marine fish and some have gained notoriety as major limiting factors in the development of marine aquaculture in various parts of the world. Among these, Vibrio spp., Aeromonas salmonicida and Photobacterium damselae susbsp. piscicida usually cause rapid mortalities as a result of hemorrhagic septicemia due to rapid tissue invasion and multiplication. Others, such as Renibacterium salmoninarum, Mycobacterium marinum and Nocardia sp., usually cause chronic, persistent infections. A

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very wide range of other bacteria and fungi can cause infections of the gills, skin, or damaged tissue—these are often regarded as opportunist pathogens. There has been extensive study on the mechanisms of pathogenicity of bacterial fish pathogens (Austin & Austin, 1999; see Chapter 17 of this volume), with toxins, adherence factors, and iron acquisition mechanisms emerging as some of the major factors in virulence. Fish are also susceptible to diseases caused by protists, commonly as a result of the activity of toxins produced during HABs. A particularly important example is the dinoflagellate P. piscicida, identified during the 1990s as a new organism responsible for massive fish kills in estuaries on the United States eastern coast (Burkholder & Glasgow, 2001). The nature of the life cycle of Pfiesteria spp. and the involvement of toxins in fish mortality have been the subject of intense controversy (Berry et al., 2002; Burkholder & Glasgow, 2002; Litaker et al., 2002). Other toxic genera, including Alexandrium, Gymnodinium, and Pseudo-nitzschia, have all been responsible for mass mortality in both wild and farmed fish (Landsberg, 2002; see Chapter 8 of this volume). Nontoxic algae can also kill fish. For example, there have been significant economic losses in salmon culture in waters off British Columbia and Washington State due to the barbed diatom Chaetoceros convolutus and the flagellate Heterosigma carterae, which interfere with gill function (Taylor, 1993). The incidence of fish mortalities from HABs has increased considerably in the last few years, largely as a result of the increased use of coastal waters for aquaculture, which contributes to the stimulation of blooms by the input of excess nutrients (waste food and excreta). Pollution of coastal waters by sewage and farm wastes also contributes to HAB formation, although the interplay of nutrient enrichment with hydrographic and other factors is complex (Anderson et al., 2002; Roelke & Buyukates, 2001). A large number of protozoa can infect wild and cultured fish. These are often free-living organisms or benign parasites, and little is known about the nutritional and environmental conditions that promote disease. For example, Paramoeba (Neoparamoeba) pemaquidenisis causes amoebic gill disease, one of the major problems in salmon culture in Tasmania (Nowak et al., 2002), but wild fish do not appear to be a significant reservoir for the pathogen (DouglasHelders et al., 2002). Diplomonad flagellates appear to be frequent harmless commensals in the gut of many fish, but under some conditions they can cause systemic infections with high mortalities (Poynton & Sterud, 2002). Myxosporeans such as Kudoa sp. can infect the muscle tissue of fish, causing softening and impairment of flesh quality (Dawson-Coates et al., 2003). Intracellular microsporidean parasites cause gill infections in many commercially important fish, including wild and cultured salmon and cod (Lom, 2002).

1.2.12. Pathogens of Marine Mammals Mass mortalities of seals, dolphins, and porpoises in European and North American waters in the late 1980s led to monitoring programs throughout the world. Use of serological and molecular diagnostic tests on tissue sampled from pinnipeds (seals and sea lions) and cetaceans (whales, dolphins, and porpoises) in mass mortality outbreaks, strandings and fishing by-catch has led to a large body of knowledge about viral pathogens of marine mammals. Viruses from nine different families have been linked to diseases of marine mammals (Table 1.1), the most important being the RNA viruses of the paramyxovirus group known as the morbilliviruses (Van Bressem et al., 1999). The virus responsible for mass mortalities (over 60% population loss in some areas) in European seals in 1988 and 2002 (Barrett et al., 2003) is closely related to canine distemper virus and the epizootic was originally thought to have been acquired from dogs, but

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the seal virus was identified as a new species (phocine distemper virus, PDV) using sequencing of viral capsid proteins (de Swart et al., 1995). The morbilliviruses isolated from diseased porpoises, dolphins, and whales are closely related genetically and antigenically and are now classed as different strains of the cetacean morbillivirus. The morbilliviruses cause pneumonialike symptoms and affect the swimming and navigation ability of infected animals. In the enzootic state, PDV and CDV probably have long-term effects on population dynamics because, like all morbilliviruses, they lead either to rapid death in the young or to life-long immunity (Van Bressem et al., 1999). Morbilliviruses require large populations in order to be sustained through the input of new susceptible hosts. Epizootics probably occur as a result of cross-infection from different species of animals or other geographic regions; migrations in response to largescale ecological changes are significant in this respect. Impairment of the immune system by environmental pollutants such as polychlorinated biphenyls, dioxins, and organochlorines has been linked to increased susceptibility to viral infections. However, controlled experiments in this area are difficult to conduct and most evidence, although persuasive, is circumstantial (Jepson et al., 1999; Ross, 2002). The most important bacterial pathogen of marine mammals is probably Brucella, strains of which have been demonstrated consistently by cultural and serological tests in many parts of the world since the 1990s (Foster et al., 2002). Members of the genus Brucella are highly contagious intracellular pathogens that also occur in cattle and sheep, where they multiply in the reproductive tract. Different species of Brucella cause abortion in particular hosts because they show tropism to specific compounds in the placenta and amniotic sac and genome analysis of Brucella spp. is leading to an understanding of these processes (Tsolis, 2002). Strains isolated from marine mammals have been designated as the new species Brucella cetaceae and B. pinnipediae (Cloeckaert et al., 2003). Brucella infection has been shown to induce abortion in dolphins (Miller et al., 1999) and the bacterium is suspected of having significant effects on fertility and population dynamics. Other bacteria thought to cause significant morbidity and mortality in marine mammals include Leptospira interrogans and related species (Stamper et al. 1998), Mycobacterium, Erysipelothrix, Salmonella, Bordetella, and Pasteurella (Higgins, 2000). In addition to primary pathogens, marine mammals (particularly stressed animals in captivity) can succumb to a variety of opportunist respiratory and skin infections by bacterial and fungal pathogens, such as pneumonia caused by Aspergillus spp. (Young et al., 1999). It should be noted that a number of infections of marine mammals can be transmitted to humans. Some well-described zoonoses exist in whale and seal hunting communities (Buck & Schroeder, 1990) and bacterial and viral infections may also be acquired from living or dead animals by research workers and aquarium staff (Higgins, 2000; Tryland, 2000). The increasing popularity of tourist activities such as swimming with dolphins could lead to a rise in transmission of zoonoses, or transfer of human pathogens to the mammals, but there is no evidence of this occurring to date.

1.2.13. Pathogens of Humans Almost all of the remaining chapters in this book deal extensively with aspects of the various types of human pathogen, so no further description of them will be provided here, apart from a few general introductory remarks. Marine microbial pathogens of humans may be considered from several different perspectives (Fig. 1.3). On the one hand, it is useful to distinguish between indigenous (or autochthonous) organisms and introduced (or allochthonous)

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Figure 1.3. Sources of human diseases in the sea, showing routes of transmission and representative pathogens.

organisms. The former are present as members of the normal constituents of plankton or are associated as colonists of marine animals or algae. The most common form of disease associated with such organisms is via the consumption of fish or shellfish containing the organisms or their products. With a few notable exceptions, discussed below, these organisms have evolved in the marine environment and their contact with, or pathogenicity for, humans can be regarded as “accidental.” Humans play no role in the normal ecology of the microbes. By contrast, many other marine-associated pathogens are bacteria, viruses, and protozoa that normally colonize humans and are excreted in the feces. These organisms are introduced to seawater via sewage contamination. Infection occurs via consumption of shellfish (or, to a lesser extent, finfish), which have accumulated the microbes, or by direct contact due to occupational or recreational exposure. A third source of human pathogens may be classed as zoonoses, that is diseases of animals such as fish and marine mammals, which are occasionally transmitted to humans. It is also convenient to classify diseases as infections or intoxications. A number of marine microorganisms (principally bacteria, dinoflagellates, and diatoms) produce toxins with a direct effect on humans. Disease from this source usually occurs when humans consume fish or shellfish in which the microbes have grown or accumulated, in which case it can be described as food-borne intoxication. Proliferation of the organism with the human body is not required and disease signs often appear within a few minutes of hours of exposure. By contrast, other pathogens (bacteria, viruses, and protozoa) associated with fish or shellfish multiply within the body after ingestion, and may be described as food-borne infections. Toxins produced

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in vivo are often important in the pathogenicity of bacteria that produce disease following infection, but their nature is usually very different from those associated with the noninfectious intoxications.

1.3. CONCLUSIONS The above discussion has highlighted a number of important themes to be considered by those studying pathogens in the sea. I have emphasized the importance of considering the wider ecological aspects of microbial life in the sea. Knowledge of the environmental factors that determine microbial abundance and activity in different habitats is vital for an understanding of the natural ecology of pathogens and the fate of introduced organisms. Comparative studies reveal many common themes, for example: (a) in the mechanisms of bacterial symbiosis and pathogenicity; (b) the importance of viruses in genetic exchange as a driving force in evolution; (c) the effect of movement of organisms between geographic areas; and (d) the effect of climatic conditions on disease occurrence. What are the most obvious gaps in our knowledge? Perhaps the most obvious is an answer to the question “are any of the Archaea pathogenic?” As noted above, it is now realized that members of the Archaea are prominent members of numerous marine ecosystems, including sediments, hydrothermal vent systems, seawater and the surfaces, and interior of animals and plants. Given the wide distribution and abundance of the Archaea, and especially their interactions with vertebrate and invertebrate hosts as symbionts or members of the gut microbiota, it is surprising that no pathogenic members of the group have been identified. In part, this may be due to our current inability to culture many Archaea in the laboratory, or it may indicate that they deploy pathogenic strategies that we do not currently recognize. With the growing amount of genomic sequence data, it is becoming possible to search for archaeal genomes for evidence of potential pathogenicity factors (Cavicchioli et al., 2003; Faguy, 2003), and the possibility of archaeal involvement in marine-related diseases of unknown etiology should certainly be considered in future investigations. A second major gap in knowledge is the role of viruses as pathogens of marine organisms. On the assumption that every cellular organism will be host to at least one type of virus, we have clearly discovered only a tiny fraction of the possible marine viruses. Increased use of molecular techniques will undoubtedly reveal many more candidates. For example, Culley et al. (2003) amplified conserved sequences of the RNA-dependent RNA polymerase gene from seawater samples, and identified four distinct groups of picorna-like viruses as persistent and widespread members of the seawater community, most of which were previously unknown. Repeated amplification of these viruses was observed over a 4-year period, suggesting that there is persistent viral production and hence infection. In the near future, such approaches should lead to isolation, full sequencing, and identification of the unknown hosts for many new viral pathogens in the sea. Study of the role of viruses in the mortality of marine life will have great influence on the development of control measures for economically important pathogens of humans and aquaculture species and ecologically important diseases of other organisms. As a final comment, it is important in future studies to bear in mind the many examples of the multifactorial nature of disease. Pathogenicity often depends on subtle interactions between several microbes, the host, and environmental factors. Growing evidence of intercellular

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communication and genetic exchange shows that microbes do not exist in isolation. For many diseases it is mistaken to look for a simple cause and effect due to a single pathogen without regard to other factors.

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2 Diversity, Sources, and Detection of Human Bacterial Pathogens in the Marine Environment Janelle R. Thompson, Luisa A. Marcelino, and Martin F. Polz 2.1. INTRODUCTION Disease outbreaks in marine organisms appear to be escalating worldwide (Harvell et al., 1999, 2002) and a growing number of human bacterial infections have been associated with recreational and commercial uses of marine resources (Tamplin, 2001). Whether these increases reflect better reporting or global trends is a subject of active research (reviewed in Harvell et al., 1999, 2002; Rose et al., 2001; Lipp et al., 2002); however, in light of heightened human dependence on marine environments for fisheries, aquaculture, waste disposal, and recreation, the potential for pathogen emergence from ocean ecosystems requires investigation. A surprising number of pathogens have been reported from marine environments and the probability of their transmission to humans is correlated to factors that affect their distribution. Both indigenous and introduced pathogens can be the cause of illness acquired from marine environments and their occurrence depends on their ecology, source, and survival. To judge the risk from introduced pathogens, levels of indicator organisms are routinely monitored at coastal sites. However, methods targeting specific pathogens are increasingly used and are the only way to judge or predict risk associated with the occurrence of indigenous pathogen populations. In this chapter, we review the recognized human pathogens that have been found in associations with marine environments (Section 2.2), the potential routes of transmission of marine pathogens to humans, including seafood consumption, seawater exposure (including marine aerosols), and marine zoonoses (Section 2.3), and we discuss the methods available to assess the public-health risks associated with marine pathogens (Sections 2.4 and 2.5).

Janelle R. Thompson, Luisa A. Marcelino, and Martin F. Polz • Department of Civil and Environmental Engineering, Massachusetts Institute of Technology, 77 Massachusetts Avenue, Cambridge, MA 02139, USA.

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2.2. DIVERSITY AND ECOLOGY Our current knowledge of the diversity and ecology of bacterial pathogens associated with marine environments stems from (i) clinical accounts of marine-acquired illnesses, (ii) disease outbreaks of known etiology in marine animals, and (iii) testing of marine environments for the presence of pathogen populations. In particular, surveys of environmental microbial communities based on 16S ribosomal RNA (rRNA) gene sequence diversity have revealed a large number of organisms closely related to human pathogens; however, the public health risk of many of these pathogen-like populations remains unknown. This is largely due to a poorly defined relationship between clinical isolates and pathogen-like populations detected in the environment because many methods used to detect environment populations do not possess high enough resolution to discriminate virulent from harmless strains. The genetic elements encoding virulence properties are not uniformly distributed among strains within a potentially pathogenic species. For marine pathogens, this has been explored in some detail in Vibrio species. Environmental populations of Vibrio are characterized by heterogeneous distributions of multiple virulence factors, combinations of which regulate the epidemic potential (e.g. Faruque et al., 1998; Karaolis et al., 1998; Chakraborty et al., 2000). Similarly, comparisons of the genomic diversity of clinical and environmental Vibrio vulnificus isolates suggest that seafood-borne human infections are established by a single highly virulent strain among coexisting genetically heterogeneous populations (Jackson et al., 1997). However, what leads to the occurrence of one strain over another remains poorly understood. Whether environmental conditions select for strains possessing human virulence factors is an area of increased research (e.g. Tamplin et al., 1996; Jackson et al., 1997; Faruque et al., 1998; Chakraborty et al., 2000). Such factors may include attachment mechanisms to organic matter, motility, secretion of lytic compounds, and the ability to grow rapidly under nutrientreplete conditions. Transfer of virulence properties between different species has been observed (Faruque et al., 1999; Boyd et al., 2000), and specific virulence factors (e.g., hemolysins, toxins, attachment pilli) may be borne on mobile genetic elements. Thus, environmental interaction may confer enhanced pathogenicity on a subset of an environmental population. In general, the marine environment may be a powerful incubator for new combinations of virulence properties due to the extremely large overall population size of bacterial populations and efficient mixing timescales. These natural phenomena may be further enhanced by human activity such as increased sewage input and ballast water transport (Ruiz et al., 2000) both of which introduce microbial species across geographical barriers.

2.2.1. Pathogenic Species The known diversity of human pathogens in the ocean continues to expand as the virulence of emerging pathogens is recognized. Pathogens associated with marine environments and their observed routes of transmission to humans are presented in Table 2.1. Of the 23 lineages currently characterized within the domain Bacteria by 16S rRNA phylogeny (Cole et al., 2003), six harbor human pathogens, and of these six, all lineages contain strains found as human and/or animal pathogens in marine environments (i.e., the Bacteroides-Flavobacterium group (Bernardet, 1998), the Spirochetes, the Gram-positive Bacteria, the Chlamydia (Johnson & Lepennec, 1995; Kent et al., 1998), the Cyanobacteria (Carmichael, 2001), and the Proteobacteria) (see also references in Table 2.1).

X

X

X

A. calcoaceticus

A. hydrophila

A. caviae

A. sobria

Acinetobacter

Aeromonas

Clostridium

C. lari∗ C. jejuni∗ C. botulinum (type E) X X X

ND

B. pseudomallei∗∗

Burkholderia

Campylobacter

X

B. maris

Brucella

ND

Speciesa

Humans

Genus

Marine animals X

X

X

Seafood X X X

X

X

X

Sea water X

X

X

X

Zoonoses

Observed routes of human infectionb

X

Aerosols

Hosts of marine disease

starvation > oxidative > acidic > thermal. However, it is difficult to assess the significance of these laboratory studies to environmental discharges. Bacterial cells that have passed through treatment plants with long retention times may be assumed to have reached a similar “post-stationary” phase, but cells discharged more directly (e.g., from boats) will presumably comprise a mixture of growth phases. The difference in sensitivities between growth phases has also been shown to be much less under natural and artificial light (Troussellier et al., 1998), and the effects of some forms of pretreatment, such as hypoosmotic shock (Gauthier et al., 1991), have been shown to actually decrease subsequent survival in seawater. In addition, preadaptation media do not appear to enhance the survival of all enteric bacteria. For example, Munro et al. (1987) found no effect on Staphylococcus aureus, Streptococcus faecalis, Pseudomonas aeruginosa, and Aeromonas hydrophila.

3.4. SUMMARY The literature reviewed in this chapter has shown that a wide range of biotic and abiotic factors influence the survival of enteric bacteria in seawater. Two issues arise from this review of these factors. First, there is a high degree of interaction between them. With the possible exception of pH and hydrostatic pressure, all of the factors discussed here interact (often in a synergistic manner) with at least one other factor, and in some cases, several others. Second, the relative importance of these factors will vary, depending on the particular marine environment. In shallow, well-mixed, clear seawater, one inactivating mechanism, solar radiation, is likely

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to dominate to the extent that it masks all other factors (Gameson 1984, 1988; Gameson and Gould, 1985; Davies-Colley et al., 1994, Sinton et al., 1994, 1999, 2002). Enteric bacteria are quickly inactivated in this environment, with fecal coliforms counts being reduced by up to five orders of magnitude over 1 day of sunlight exposure (Fig. 3.3). An example at the other extreme would be a sewage discharge from a deeply submerged outfall into a nutrient-rich, productive, low clarity water, with the effluent trapped under a thermocline, and thus receiving minimal sunlight exposure. Although estuarine waters are rarely very deep, a semi-saline water could be added to this scenario. As discussed in Section 3.1.4, low clarity seawaters can reduce the overall penetration of sunlight by two orders of magnitude, particularly the most biocidal wavelengths. Thus, in low sunlight marine environments, factors such as grazing and predation by marine microbiota, particle association, sedimentation, lower salinity, higher nutrient concentrations, and possibly even temperature differences may begin to have a measurable effect. Although a considerable body of information exists on biotic and abiotic factors affecting the survival of enteric indicator bacteria seawater, significant knowledge gaps remain. Sunlight inactivation rates of bacterial pathogens such as Campylobacter spp. and Salmonella spp. do not appear to have been quantified. Further research is also needed into the operation of bacterial photorepair mechanisms in seawater, the effects of prior sunlight exposure during wastewater treatment on the survival of different enteric bacteria, and the relationships between particle association, sedimentation, and resuspension.

ACKNOWLEDGMENTS The author wishes to thank Dr Rob Davies-Colley, National Institute of Water and Atmospheric Research, New Zealand, for providing the data on sunlight attenuation on seawater, and for his thorough review of this chapter. Valuable comments were also received from the book editors. Most of the figures were reproduced with the permission of Applied and Environmental Microbiology. This review was funded from the New Zealand Public Good Science Fund, administered by the New Zealand Foundation for Research, Science and Technology.

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4 Survival of Enteric Bacteria in Seawater: Molecular Aspects Yael Rozen and Shimshon Belkin 4.1. INTRODUCTION The factors controlling the survival of enteric bacteria in the marine environment have intrigued scientists for decades, and numerous studies have explored the fate of Escherichia coli and other enteric bacteria following their exposure to seawater. Most of these studies were driven by public health concerns, while others attempted to expand our basic understanding of bacterial responses to this complex environmental stress. In many of these studies, die-off rates based on colony formation capabilities were often the main parameter used to characterize the bacterial responses, under a variety of biotic and abiotic test conditions. Only in recent years has there been an increasing number of reports attempting to assay additional viability criteria, as well as preliminarily explore the molecular mechanisms involved and their genetic regulation. In the latter group, several studies focused on the role of specific genes in seawater survival (Gauthier et al., 1990, 1992a; 1993; Gourmelon, 1995; Gourmelon et al., 1997; Hengge-Aronis, 2000; Loewen et al., 1998; Munro et al., 1989, 1994, 1995; Rozen, 2001; Rozen, & Belkin, 2001; Rozen et al., 2001; Troussellier et al., 1998). In a more global approach, an E. coli promoter-reporter fusion library and DNA microarray chips were used to assay gene induction in low-nutrients seawater and to reveal RNA pattern following seawater exposure (Rozen et al., 2002). In this chapter we review these attempts to reveal the molecular mechanisms involved in determining survival and sensitivity of enteric bacteria in seawater.

4.2. GENES INVOLVED IN E. COLI SURVIVAL IN SEAWATER The search for genes that may play a role in the survival of bacteria such as E. coli was mostly based on exposure of mutants to natural or artificial seawater and a comparison of their seawater survival to their wild types. Although a relatively large number of genes were screened in this manner (66, and Table 4.1), a surprisingly small number of them were revealed as crucial to seawater survival. Yael Rozen • Technion—Israel Institute of Technology, Technion City, Haifa 32000, Israel. Shimshon Belkin • Institute of Life Sciences, Hebrew University of Jerusalem, Jerusalem 91904, Israel.

Oceans and Health: Pathogens in the Marine Environment. Edited by Belkin and Colwell, Springer, New York, 2005.

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Table 4.1. Effect of specific mutations on E. coli seawater survival.

Strain

Relevant mutation

Survival in the dark1

FRAG-5 (Epstein, & Kim, 1971) TK110 (Epstein, & Kim, 1971) TK121 (Epstein, & Kim, 1971) TK133 (Epstein, & Kim, 1971) TK142 (Epstein, & Kim, 1971) TK2205 (Munro et al., 1989) TK401 (Epstein, & Kim, 1971) TK2240 (Munro et al., 1989) PLB3260 (Benson, & Decloux, 1985) PLB3261 (Benson, & Decloux, 1985) SG477 (Garrett et al., 1983) SBM26 (Gauthier et al., 1992a) SG480 (Garrett et al., 1983) UE14 (Boos et al., 1987) RK33Z (Karpel et al., 1991) TA15D2 (Dover, & Padan, 2001) JJ2 (Jung et al., 1989)

kdpABC kdpABC, kefB (trkB) kdpABC, kefC (trkC) kdpABC, trkA kdpABC, trkE kdpABC, trkA, trkD, kdpABC, trkA, trkD trkA, trkD envZ (ompF) envZ (ompC) envZ (ompC) envZ (ompF, ompC) envZ (ompF, ompC) treA nhaA hns osmB

FF4169 (Giaever et al., 1988) RO22 (Hengge-Aronis et al., 1991) EF047 (May et al., 1986) RH90 (Lange, & Hengge-Aronis, 1991) QC2410 (Dukan, & Touati, 1996) UM122 (Loewen, & Triggs, 1984) CA8306 (Kumar, 1976) CA8445 (Sabourin, & Beckwith, 1975) CA8445-1 (Sabourin, & Beckwith, 1975) CF1652 (Xiao et al., 1991) CF1693 (Xiao et al., 1991) QC2408 (Dukan, & Touati, 1996) QC2445 (Touati et al., 1995) UM120 (Loewen, & Triggs, 1984) QC2476 (Dukan, & Touati, 1996) QC2463 (Dukan, & Touati, 1996) QC2473 (Dukan et al., 1999) QC2474 (Dukan et al., 1999) QC772 (Carlioz, & Touati, 1986) QC773 (Carlioz, & Touati, 1986) QC2472 (Dukan, & Touati, 1996) QC774 (Carlioz, & Touati, 1986) QC2589 (Dukan et al., 1999) QC2413 (Dukan, & Touati, 1996) SASX41B (Gourmelon, 1995)

otsA otsAB proP, proU rpoS rpoS rpoS cyaA cyaA, crp, rpsL crp, rpsL relA relA, spoT dps fur katE katE−, katG katE, katG, dps soxR soxS sodA sodB sodA, sodB sodA, sodB sodA, sodB, zwf oxyR hemA

− 7,8 −8 −8 −8 −8 −7 −8 −7 +9 +9 −8 +9 −8 −8 −8 −8 Protective e effect 7 +7

HB101 (Gourmelon, 1995) DM803 (Mount et al., 1972) QC227 (Gourmelon, 1995) BW295 (Gourmelon, 1995)

recA LexA (SOS repressed) fpg, uvrA xthA

−7 + 11,12 +8 −8

−8 −8 + 12 + 12 −8 −8 − −8 −8 −8 8

−8 −8

−8

Survival in the light2

Survival in the dark after pre-adaptation − 3,7

+ 3,7 + 3,7

− 10 + 10,13 + 10,13

+ 10,13 − 14 + 10 + 10 − 13,14 − 13,14 + 13,14 + 10 Protective effect14 − 13,14 − 13,14 −13,14

+ 3,7 + 4,7 [SB1] + 5,11

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Table 4.1. (Continued )

Strain CAG9333 (Kusukawa, & Yura, 1988) AB1899 (Howard-Flanders et al., 1964) CAG22188 (Rouviere et al., 1995) TN3151 (Farewell et al., 1996) CLG302 (Gauthier et al., 1990) W2961 (Gauthier et al., 1990) K10 (Willsky et al., 1973) C101a (Willsky et al., 1973) C75a (Garen, & Garen, 1963) C75b (Garen, & Garen, 1963) C112a (Garen, & Garen, 1963)

Relevant mutation rpoH (groE constitutive) lon rpoE uspA phoA phoE pit pit, pstA pit, phoR pit, phoS pit, phoT

Survival in the dark1

Survival in the light2

Survival in the dark after pre-adaptation

−8 −8 −8 −8 − 15

+ 6,15 + 6,16 − 6,16 + 6,16 + 6,16 + 6,16 + 6,16

+

Significant effect of the mutation;—no effect; no sign: mutation not tested. Survival experiments—Rozen and Belkin (2001) and Munro et al. (1995): stationary phase cells were used. In the other experiments (Gauthier et al., 1992a; Gauthier et al., 1990; Munro et al., 1989) logarithmic phase cells were used. 2. Seawater survival in the light—Experiments were in filter-sterilized artificial seawater (Instant Ocean), 20 C, 180 rpm, artificial visible light mainly >400 nm, (0.9 mmol photons.m−2 .s−1 ) (Gourmelon, 1995; Gourmelon et al., 1997; Troussellier et al., 1998). 3. Previously grown in minimal medium supplemented with NaCl (Munro et al., 1989). 4. Previously grown in minimal medium supplemented with glycine betaine (Munro et al., 1989). 5. Previously exposed for 1 hour at 37 C to oxidative (H2 O2 , 1mM), acid (pH 5.1), osmotic (0.5M NaCl), starvation (a nutrient deficient buffer) or thermal stress (48 C) (Munro et al., 1994). 6. Previously grown in phosphate-limited minimal medium (0.1 mM K2 HPO4 ) (Gauthier et al., 1990) or (0.1 mM Na2 HPO4 ) (Gauthier et al., 1993). Additional references: 7. (Munro et al., 1989), 8. (Rozen, & Belkin, 2001), 9. (Gauthier et al., 1992a), 10. (Gourmelon et al., 1997), 11. (Munro et al., 1994), 12. (Munro et al., 1995), 13. (Troussellier et al., 1998), 14. (Gourmelon, 1995), 15. (Gauthier et al., 1990), 16. (Gauthier et al., 1993). Modified from Rozen and Belkin (2001). [SB1] Effect in seawater was recorded when previously grown in the presence of GBT 1.

In fact, only six mutations were shown to have a significant effect when experiments were performed in the dark. The most dominant among them was rpoS (Gourmelon, 1995; Gourmelon et al., 1997; Munro et al., 1994, 1995; Troussellier et al., 1998). The other genes in which mutations were reported to be significant were otsA (Munro et al., 1989), relA, spoT (but only on top of a relA deletion (Munro et al., 1995)), and one or both membrane porins ompC and ompF (Gauthier et al., 1992a). An rpoS mutation was shown to decrease seawater survival (CFU) of stationary phase E. coli cells by approximately 3 logs over 8 days (Munro et al., 1995). The RpoS (σ s ) transcription factor controls the expression of a large number of genes involved in cellular responses to a diverse number of stresses, including starvation, osmotic stress, acid shock, cold shock, heat shock, oxidative damage, and transition to stationary phase. A list of over 50 genes under the control of rpoS has been compiled (Loewen et al., 1998). It normally acts as a positive regulator, but some genes appear to be under its negative control. The synthesis and accumulation of σ s were shown to be controlled at the transcription, translation, holoenzyme complex formation, and proteolysis levels. Transcriptional control of rpoS involves guanosine 3
,5
bispyrophosphate (ppGpp) and polyphosphate as positive regulators, and the cAMP receptor protein-cAMP complex (CRP-cAMP) as a negative one (Loewen et al., 1998). Munro et al. (1995) demonstrated that in addition to rpoS, a relA deletion and the double relA spoT mutation also affected seawater survival of E. coli. The absence of these two genes

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from stationary phase E. coli was shown to cause an approximately 4 logs decrease in E. coli colony formation ability over 8 days. RelA and SpoT are enzymes, ppGpp synthetase I and II respectively, that are responsible for the synthesis of 3
,5
-bispyrophosphate (ppGpp) (Cashel et al., 1996). Among the multiple effects caused by the increase in intracellular ppGpp concentration (Cashel et al., 1996) is an enhancement of rpoS expression (Gentry et al., 1993). The hypothesis that the intracellular level of ppGpp could affect seawater survival due to lower rpoS expression was suggested (Munro et al., 1995), but was not experimentally tested. An additional gene shown to be important in determining E. coli’s sensitivity to seawater was otsA (Munro et al., 1989); within an 8-day experiment, colony formation ability of the mutant (E. coli FF4169) decreased 10–100 fold compared to the wild type. The otsA gene codes for trehalose–6-phosphate synthase, an enzyme that catalyzes the synthesis of the osmoprotectant trehalose; it is transcriptionally activated both by osmotic stress (Giaever et al., 1988) and by growth into the stationary phase (Hengge-Aronis et al., 1991). Not surprisingly, otsA is RpoS controlled (Hengge-Aronis et al., 1991; Kaasen et al., 1992), and it is not unlikely that at least part of the rpoS deleterious effect on E. coli seawater survival is due to the absence of trehalose. Mutants lacking OmpC, OmpF, or both also seemed to be impaired in their seawater survival. The relative significance of the two porins depended upon pregrowth conditions: ompF was more dominant following exponential growth in low osmolarity, while ompC appeared to be more important for survival following exponential growth in high osmolarity (Gauthier et al., 1992a). It should be noted that standard deviations in these experiments were large, and not all differences appeared significant. In a different experiment (Rozen, & Belkin, 2001), other E. coli strains lacking OmpC or both OmpC and OmpF were pregrown to stationary phase in a minimal medium. No effect on survival of either mutation was observed during a 7-day seawater exposure (Table 4.1). Other genes were also implicated in the ability of E. coli to form colonies on solid media, but only in conjunction with additional stress factors or following specific growth conditions. One group of genes shown to be important for seawater survival when previously grown in a low phosphate (0.1 mM Na2 HPO4 ) medium is connected with phosphate regulation. Mutation in the porin phoE, in the regulatory gene phoR, and in genes related to the high-affinity phosphate specific transport system pstA, phoS, and phoT decreased survival 10–1500 fold after 6 days in filtered sterilized seawater (Gauthier et al., 1993). Furthermore, an E. coli strain totally deprived of alkaline phosphatase activity due to a phoA mutation was found to have increased seawater sensitivity when previously grown in a low phosphate (0.1 mM K2 HPO4 ) medium (Gauthier et al., 1993). Numerous genes involved in protection against reactive oxygen species (rpoS, katE katG, dps, sodA sodB, and oxyR) proved to be important when the cells were exposed to visible light (Gourmelon, 1995; Gourmelon et al., 1997) (see Section 4.1) The absence of the two genes encoding for glycine betaine transport system, proP and proU, was found to increase E. coli MC4100 seawater sensitivity only when previously exposed for 1 h to high osmolarity and glycine betaine (0.1 mM) (Munro et al., 1989). Similarly, mutations in two potassium transport systems, kdpABC (osmotically inducible high-affinity transport, E. coli TK2240) and trkA (constitutive low-affinity transport, E. coli TK2205), were found to have a deleterious effect on E. coli (MC4100) survival in seawater only when previously grown in a minimal medium at high osmolarity (0.5 M NaCl) (Munro et al., 1989). The absence of one gene, induced by high osmolarity, but with an unknown function— osmB—was found to have a surprisingly protective effect on seawater survival compared to its

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Molecular Aspects of Survival

Table 4.2. Effect of specific mutations on E. coli growtha in nutrient-supplemented seawater. Strain

Relevant mutation

FRAG-5 (Epstein, & Kim, 1971) TK110 (Epstein, & Kim, 1971) TK121 (Epstein, & Kim, 1971) TK133 (Epstein, & Kim, 1971) TK142 (Epstein, & Kim, 1971) TK401 (Epstein, & Kim, 1971) SG477 (Garrett et al., 1983) SG480 (Garrett et al., 1983) UE14 (Boos et al., 1987) RK33Z (Karpel et al., 1991) TA15D2 (Dover, & Padan, 2001) QC2410 (Dukan, & Touati, 1996) CA8306 (Kumar, 1976) CA8445 (Sabourin, & Beckwith, 1975) CA8445-1 (Sabourin, & Beckwith, 1975) QC2408 (Dukan, & Touati, 1996) QC2445 (Touati et al., 1995) QC2476 (Dukan, & Touati, 1996) QC2463 (Dukan, & Touati, 1996) QC2473 (Dukan et al., 1999) QC2474 (Dukan et al., 1999) QC2472 (Dukan, & Touati, 1996) QC2589 (Dukan et al., 1999) DM803 (Mount et al., 1972) CAG9333 (Kusukawa, & Yura, 1988) AB1899 (Howard-Flanders et al., 1964) CAG22188 (Rouviere et al., 1995) TN3151 (Farewell et al., 1996)

kdpABC kdpABC, kefB (trkB) kdpABC, kefC (trkC) kdpABC, trkA kdpABC, trkE kdpABC, trkA, trkD envZ (ompC) envZ (ompF, ompC) treA nhaA hns rpoS cyaA cyaA, crp, rpsL crp, rpsL dps fur katE-, katG katE, katG, dps soxR soxS sodA, sodB sodA, sodB, zwf LexA (SOS repressed) rpoH (groE constitutive) lon rpoE uspA

a

Growth − − − + + − − + − + + + − − − − − − − − − + + − − − + −

Overnight cultures were diluted 100-fold in filtered (0.22 µm) Luria Bertani (LB) broth (Miller, 1972) and seawaterbased LB. Growth (37◦ C, 200 rpm) was followed by culture turbidity for 9 hours. Experiments were repeated at least 3 times, and the mutation was noted as + when the wild type strain grew significantly better (relative growth yield at least 2 fold higher) than the mutated strain. Relative growth yield was defined as the ratio of the yield in LB to that in seawater-based LB. Modified from Rozen and Belkin (2001).

parental strain (Munro et al., 1989). No explanation was provided for this phenomenon. The same effect was observed for an E. coli strain mutated in hemA, when incubated in seawater under visible light. In this case, a decrease in intracellular photosensitizer concentrations was suggested (Gourmelon, 1995). The marine environment is notoriously poor in nutrients; consequently, the studies quoted above were carried out by incubating the bacteria in nutrient-poor seawater. Nevertheless, enteric bacteria may encounter relatively nutrient-rich conditions close to marine outfall or wastewater disposal areas. Several mutations that did not have an effect on survival in nutrientfree media turned out to affect essential genes when the strains carrying them were forced to grow in nutrient-amended seawater (Rozen, & Belkin, 2001). The data in Table 4.2 include the results of such growth experiments, each mutant compared to its own wild-type parental strain. An effect was noted when the deleterious influence of the mutation on growth in seawater-based LB was significantly higher than its effect on growth in regular LB. Among the genes with significant effects were some involved in osmoregulatory processes (kdpABC trkA, kdpABC trkE, ompF ompC), defenses against oxidative stress (sodA sodB), the sigma factor rpoE, the

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gene encoding to the histone-like protein H-NS (hns), and the sodium proton antiporter (nhaA). The effect of the latter mutation was the most dramatic: no growth was observed at all in the mutant. The mutation in rpoS had the expected effect in both incubation conditions.

4.3. GENE INDUCTION UPON EXPOSURE OF E. COLI TO SEAWATER A more global strategy in the search for genes playing a role in seawater survival was demonstrated by Rozen et al. (2001) who screened a 687-member E. coli promoter:: luxCDABE fusion library with Photorhabdus luminescens luxCDABE as a reporter (Van-Dyk et al., 2001). Each strain harbored a plasmid-borne lux fusion of a different promoter region of E. coli K12 in host strain E. coli DPD1675 (Van-Dyk, & Rosson, 1998; Van-Dyk et al., 2001), and all were treated by a shift from a rich environment (LB) to a combination of nutrient scarcity, osmotic stress, and elevated pH (Rozen et al., 2001). These unfavorable conditions led to a reduction in the light production of the majority (ca. 90%) of the strains, while in the expression of approximately 7% of the promoters there was no significant change. In contrast, bioluminescence of approximately 4% of the tested promoters increased over 1.5 fold upon exposure to ASW. Of these, the luminescence of 22 fusions was reproducibly increased at least twofold in ASW versus LB (Rozen et al., 2001)—osmY, ldcC, treA, otsA, poxB, yciG, ybhP, yohF, yohC, ybdK, yeaG, yhdW, prpR, ytrF, ykgC, phrB, yjhT, ykgD, cysI, pstS, ygjL, and yjbG. Only nine of these genes have previously been assigned a function (Rozen et al., 2001). As has already been indicated, seawater survival of E. coli is greatly affected by σ s (Munro et al., 1994, 1995). This observation received a dramatic confirmation when promoter activity from all 22 ASW-induced plasmids was tested in an isogenic pair of E. coli strains, one of them an rpoS mutant. Seventeen out of 22 promoters were found to be under rpoS positive control and one (yjbG) appeared to be negatively regulated by this protein (Rozen et al., 2001). When the combined effect of ASW was broken into its different environmental factors, using a set of distinct media imitating salinity, nutrient limitation, alkaline pH, and their combinations (Rozen et al., 2001), it appeared that salinity or osmolarity were involved in only four cases, among them the expected osmY, as well as the previously “unassigned” genes yciG, ybhP, and yjbG. For the other promoters, the most important environmental factor appeared to be nutrient limitation; this observation included otsA and treA, both known to be osmotically regulated, which seemed in this system to be induced by nutrient limitation rather than by salinity (Rozen et al., 2001).

4.4. RNA PATTERN OF SEAWATER-INCUBATED E. COLI A different attempt to reveal the overall pattern of responses induced by the shift to the hostile marine environment was made using whole-genome DNA array technology (Rozen et al., 2002). Using the DNA microarray method (Wei et al., 2001), Rozen et al. (2002) identified E. coli genes the expression of which was changed following 20 h of seawater exposure, compared to the same cells growing in MOPS-glucose medium. It was assumed that by then “emergency operations” will be over and that only genes involved in continuous maintenance of viability will be transcribed. Following treatment, the expression of most (ca. 3000) of the 4228 ORFs

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on the chip remained unchanged. Out of the rest, the relative expression of about one-fourth of the genes (937) increased in seawater while that of a smaller fraction (320) decreased (Rozen et al., 2002). An analysis of the expression of functional gene groups (Serres & Riley, 2000) revealed several clear expression patterns that reflect major physiological changes in E. coli’s gene transcription profile. Among the gene groups the expression of which was down-regulated, most notable were genes involved in cell division and in nucleotide biosynthesis. In contrast, an increase was observed in the expression of genes the products of which are involved in either carbon compound degradation or in respiration, signifying the dearth of carbon and energy sources under such conditions. An additional potential that was built within this period was motility, equipping the cells either for a chemotactic search for nutrition or for flight from the inhospitable conditions imposed upon them.

4.5. E. COLI PROTEIN SYNTHESIS IN SEAWATER E. coli was shown to synthesize specific sets of proteins in response to individual stress conditions involved in seawater stress, such as carbon, phosphate, or nitrogen starvation (Groat et al., 1986), osmotic stress (Botsford, 1990; Clark, & Parker, 1984), cold shock (Jones et al., 1987), elevated hydrostatic pressure (Welchet al., 1993), and oxidative stress (Walkup & Kogoma, 1989). Starvation was shown to induce cross protection against osmotic stress (Jenkins et al., 1990) and H2 O2 (Jenkins et al., 1988), attributed in both cases to proteins induced under both conditions. Protein synthesis de novo was shown to be important for E. coli survival under carbon starvation conditions by using the protein synthesis inhibitor chloramphenicol (Reeve et al., 1984a). The earlier during starvation chloramphenicol was added, and the higher its concentration, the greater was its deleterious effect on cell viability (Reeve et al., 1984a). Protein degradation was also found to play a role in the survival of carbon-starved E. coli (Reeve et al., 1984b), increasing when the cells were subjected to nutrient starvation (Goldberg & St John, 1976). A carbon-starved E. coli mutant lacking five distinct peptidase activities lost viability faster than its parental strain (Reeve et al., 1984b). In a carbon-deficient medium this mutant was also defective in its protein synthesis ability, possibly due to unavailability of free amino acids in the starved mutant cells (Reeve et al., 1984b). Only one study addressed the question of enteric bacterial proteins newly synthesized in the marine environment (Gauthier et al., 1989b). The overall protein content of E. coli cells during incubation in sterile seawater was observed to gradually decrease by approximately 40% after 14 days of seawater incubation, while some new cytoplasmic membrane proteins were observed (Gauthier et al., 1989b). The development of these changes was considerably slower in cells that had previously been grown in marine broth, suggesting a protective effect of osmotic regulation mechanisms. In a set of experiments designed to directly assay protein synthesis upon a shift to a seawater environment, 3 H-isoleucine incorporation was monitored in E. coli following exposure to seawater. Incorporation of 3 H-isoleucine into proteins was reduced to undetectable levels immediately after the shift, but resumed after approximately 2 h (Fig. 4.1A). Incorporation rate continued to increase for about 24 h, and was maintained at that level for at least 2 weeks (Fig. 4.1B) (Rozen, 2001).

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A.

B. 1.6×104 1.4×104

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Figure 4.1. 3 H-isoleucine incorporation into proteins of stationary-phase E. coli cells following exposure to seawater. E. coli MG1655 was grown in a MOPS-glucose (0.2%) medium to stationary phase, washed and re-suspended in filtered seawater (Tel Aviv beach) to a density of about 108 cells/ml, and incubated at room temperature with 14 Hlabeled isoleucine. Results are average and standard deviations of 2–3 different experiments (Rozen, 2001), A time course of incorporation and B incorporation rate as a function of exposure duration.

The significance of this activity to cell viability was addressed by following colony formation of cells exposed to seawater in the absence or presence of the protein synthesis inhibitor chloramphenicol (Rozen, 2001). As previously shown (Gauthier et al., 1992b), stationary phase cells survived much better in seawater than exponentially growing ones; surprisingly, however, chloramphenicol had no significant effect on the viability of both cell types (Fig. 4.2A). As shown in Figure 4.2B, chloramphenicol was still active in protein synthesis inhibition after

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B. 40000

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Logarithmic 0.1 0

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35000 30000 25000 20000

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Figure 4.2. A Survival of E. coli in seawater in the presence and absence of the protein synthesis inhibitor chloramphenicol (100 µg/ml). B 3 H-isoleucine incorporation into proteins of logarithmic and stationary phase E. coli cells with or without chloramphenicol following 6 days incubation in seawater. Cell preparation as in Figure 4.1. Results are averages and standard deviations of six different experiments. 3 H-isoleucine incorporation tests of up to 1 h were carried out after 1, 6, and 15 days to verify chloramphenicol protein synthesis inhibition after these periods of seawater incubation.

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6 days in seawater (Rozen, 2001), indicating that it did not lose its potency during the experiment. While it is therefore evident that seawater exposure of E. coli is accompanied by protein synthesis, there are no clear indications at the moment as to the role the newly synthesized protein may play in survival.

4.6. DISCUSSION 4.6.1. The Importance of RpoS A recurring point arising from the accumulated data is that seawater incubation conditions, as well as previous growth history, have a major influence on subsequent survival in the hostile marine environment. This clearly indicates the existence of an adaptation potential: at least some of the risks inherent in seawater exposure can be handled better if the cells can first mobilize the necessary defense circuits. The molecular data available to date clearly point at possibly the most significant among such adaptive systems: the rpoS regulon. The identification of genes the activity of which is significant to seawater survival has yielded a very short list of mutations affecting survival of E. coli in sterile seawater. Among the mutations with the most dominant effect were mutations in relA spoT (Munro et al., 1995) and rpoS (Gourmelon, 1995; Gourmelon et al., 1997; Munro et al., 1994, 1995; Troussellier et al., 1998). RelA and SpoT are enzymes, ppGpp synthetase I and II respectively, that are responsible for the synthesis of guanosine 3
,5
-bispyrophosphate (ppGpp) (Cashel et al., 1996). Among the multiple effects attributed to the increase in intracellular ppGpp concentration (Cashel et al., 1996) is an enhancement of rpoS expression (Gentry et al., 1993). RpoS (σ s ) is a transcription factor that controls the expression of a large number of genes, among them otsA (HenggeAronis et al., 1991), a mutation in which was also shown to enhance seawater sensitivity (Munro et al., 1989). RpoS is involved in cellular responses to a diverse number of stresses, including starvation, osmotic stress, acid shock, cold shock, heat shock, oxidative damage, and transition to stationary phase (Loewen et al., 1998). Transcriptional control of rpoS involves ppGpp and polyphosphate as positive regulators, and the CRP-cAMP as a negative one (Loewen et al., 1998). The hypothesis that the intracellular level of ppGpp could affect seawater survival due to lower rpoS expression was suggested (Munro et al., 1995), but was not experimentally tested. The small number of genes that were shown to be essential for seawater survival probably indicates a large degree of functional overlap in the activities of the other genes the mutations of which were tested. In view of this, it is of significance that all “positive” mutations that were not in the rpoS gene itself were closely linked to it. Apparently the importance of this gene to survival in the marine environment is such that there is no efficient molecular backup for it. Further proof for the importance of the rpoS regulon was provided by the use of an extensive promoter::lux fusion library: of the 22 identified ASW-induced promoters, 17 were subject to positive RpoS regulation. Additional support to the central role rpoS plays in determining survival is provided by the observation that previous growth history has a major influence on subsequent survival of enteric bacteria in the marine environment (Garcia-Lara et al., 1993; Gauthier et al., 1987, 1989a, 1989b, 1990, 1991, 1992b; Munro et al., 1987a, 1987b, 1989, 1994; Troussellier et al., 1998). Cross protection against seawater stress endowed by

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oxidative, acidic, thermal, and nutritional stresses was rpoS dependent (Munro et al., 1994). Furthermore, it has been experimentally demonstrated that stationary phase cells (in which rpoS regulated genes are induced) are far less sensitive to seawater than logarithmically growing cells (Gauthier et al., 1992b). In view of these data it is somewhat surprising that analysis of the DNA microarray data (Rozen et al., 2002) has revealed no clear increase in expression of genes known to be positively controlled by rpoS (Hengge-Aronis, 1996; Rozen et al., 2001; Van-Dyk et al.,1998). It is possible that most of these rpoS-regulated effects were more apparent during the first few hours of exposure, and were of lesser significance 20 h later.

4.6.2. Nutrient Limitation versus Osmotic Stress Of the different environmental stress factors confronted by an enteric bacterium released into the sea, it would seem reasonable to expect that osmotic stress would be the most important. Indeed, studies focusing on the different environmental parameters affecting seawater survival show the deleterious effect of salinity on seawater survival (Anderson et al., 1979; Carlucci, & Pramer, 1960b) as well as the importance of preadaptation to osmotic stress (Gauthier et al., 1987; Munro et al., 1987b, 1989). Nevertheless, most of the genes we have found to be activated upon seawater exposure are related to lack of nutrients (Rozen et al., 2002). It has already been shown that in the presence of sufficient nutrients, E. coli can grow in seawater almost as well as it does in rich laboratory media. This points at the importance of nutrients availability, rather than salinity, as limiting enteric bacteria in the sea. A support for this was provided by the results of the promoter::lux library screen. Out of the 22 identified seawater-induced promoters all but one seemed to respond primarily to nutrient limitation, while only 4 were induced by salt stress (Rozen et al., 2001). A very similar trend was apparent in the pattern emerging from the gene expression arrays. Following 20 h in seawater, the main efforts of the cells appeared to be directed toward obtaining carbon and energy. Interestingly, induction of enzymes essential for the degradation of fatty acids, amino acids, and other small carbon compounds was much more pronounced than pathways for macromolecule degradation. It is possible that at this stage the efforts are still directed toward obtaining soluble molecules from an external source rather than toward the consumption of internal reserves (Rozen et al., 2002). At the same time, no clear pattern could be discerned among the group of genes responsible for maintaining intracellular osmotic stress, although some of them were clearly induced. Thus, it seems as if after 20 h in seawater the major challenge faced by the cells is similar to that after a few hours only: acquisition of nutrients and energy and not maintenance of osmotic equilibrium.

4.6.3. The Role of Protein Synthesis in Seawater Another attempt to understand cellular processes in E. coli cells in seawater involved a study of protein synthesis. Protein synthesis was found to continue for at least 2 weeks in cells incubated in sterile seawater (Rozen, 2001). As nutrients were not added to the medium, protein synthesis de novo under such conditions had to be accompanied by protein degradation in order to supply amino acids for the new proteins synthesized. Indeed, an increase in protein degradation under starvation conditions was shown (Goldberg & St John, 1976; Reeve et al., 1984b), as well as

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an increased viability loss under carbon-starved E. coli in a peptidase-deficient mutant (Reeve et al., 1984b). Protein synthesis de novo was previously shown to be important for E. coli survival under carbon starvation conditions (Reeve et al., 1984a). The importance of protein degradation was also attributed to reduced protein synthesis due to lack of free amino acids (Reeve et al., 1984b). As all of these data indicate a potentially important role of protein synthesis for seawater survival; it was surprising that inhibition of protein synthesis had no significant effect on survival (Rozen, 2001). Additional support for the potentially minor role of de novo protein synthesis under field conditions is the enhanced stability of E. coli at lower temperatures (Carlucci & Pramer, 1960a; Lessard & Sieburth, 1983; Vasconcelos & Swartz, 1976), in which protein synthesis is partially or completely diminished. Thus, E. coli synthesizes proteins in seawater but this activity does not contribute to its survival compared to cells in which protein synthesis in inhibited. The role of these proteins is not yet clear, though a few explanations may be put forward in line with the overall pattern of gene expression that arises from this review. It is possible, for example, that at least some of the newly synthesized proteins may help develop the potential for the exploitation of low-nutrients concentrations if they become available. Such phenomena are not unknown. For instance, it was shown that in carbon-limited continuous cultures, cells exhibit a higher activity of several catabolic enzymes even in the absence of the appropriate carbon sources and are able to utilize a wide range of sugars without an induction lag (Harder & Dijkhuizen, 1983; Lendenmann & Egli, 1995). Another hypothesis regarding the potential role of the newly synthesized proteins may relate to an enhancement of chemotaxis and motility capabilities. These may be highly important in a natural environment, where different niches exist and in which nutrient gradients may be sensed. In the uniform experimental systems of the present study, however, neither motility nor chemotaxis would confer any advantage; survival was thus not affected by the inhibition of the synthesis of the relevant proteins.

4.6.4. Future Directions and Public Health Implications The use of new methodologies such as DNA arrays, in conjunction with traditional physiological, biochemical, and molecular approaches, opens a way for a broader spectrum of questions and allows a more detailed research of bacterial molecular responses to environmental stresses. For example, the ability to monitor the expression of the entire genome of an organism as a function of time, from the moment of exposure to complete loss of culturability and/or viability, may clarify some of the mechanisms determining survival under stress. While the list of genes the activity of which is essential for seawater survival is bound to expand, the significance of some will not be immediately apparent. Quite clearly, the understanding of the roles of these genes and their interactions will require the use of a broad range of viability criteria. As demonstrated here, the use of the bacterial lux operon as a reporter system allowed early sensitive detection of gene induction. To allow the use of this technology in field studies, the application of additional reporters should be considered. Green fluorescence protein was demonstrated as a useful reporter system in natural waters, allowing detection of gene expression in aquatic environments for several days (Espinosa-Urgal & Kolter, 1998, 1999) without the necessity of adding external nutrients. With the increasing availability of fluorescent reporter proteins, it should be possible to design diverse molecular probes for enteric bacterial presence or activity in marine water.

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Finally, one should remember that even though E. coli serves as a convenient model organism, it is important that true pathogenic bacteria are studied as well, with a greater focus on infectivity bioassays in addition to the various viability parameters.

ACKNOWLEDGMENTS The use of many E. coli strains from the collection of T. K. Van Dyk and R. A. LaRossa (Dupont, Wilmington, DE, USA), S. Dukan (CNRS/INSU UMR, Marseille, France), and M. Berlyn at CGSC (E.coli Genetic Stock Center, Yale University, New Haven, CT, USA) is gratefully acknowledged. We also thank Karl Forchhammer (Institut f¨ur Mikrobiologie und Molekularbiologie der Justus-Liebig-Universit¨at Giessen, Heinrich-Buff-Ring 26–32, 35392 Giessen, Germany) for the hospitality and help in the protein studies. This study was supported by The Israel Science foundation grant number 113-00-1 and by the Israeli Ministry of Science and the Arts. Yael Rozen was a Rieger-JNF fellow.

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Touati, D., Jacques, M., Tardat, B., Bouchard, L., & Despied, S. (1995). Lethal oxidative damage and mutagenesis are generated by iron in –fur mutants of Escherichia coli: Protective role of superoxide dismutase. J Bacteriol 177, 2305–2314. Troussellier, M., Bonnefont, J.L., Courties, C., Derrien, A., Dupray, E., Gauthier, M., Gourmelon, M., Joux, F., Lebaron, P., Martin, Y., & Pommepuy, M. (1998). Responses of enteric bacteria to environmental stresses in seawater. Oceanologica Acta 21, 965–981. Van-Dyk, T.K., Ayers, B.L., Morgan, R.W., & LaRossa, R.A. (1998). Constricted flux through the branched-chain amino acid biosynthetic enzyme acetolactate synthase triggers elevated expression of genes regulated by rpoS and internal acidification. J Bacteriol 180, 785–792. Van-Dyk, T.K., & Rosson, R.A. (1998). Photorhabdus luminescens luxCDABE promoter probe vector, vol. 102. In R.A. LaRossa (ed.), Methods in Molecular Biology, Bioluminescence Methods and Protocols. Humana Press Inc., Totowa, NJ, pp. 85–95 Van-Dyk, T.K., Wei, Y., Hanafey, M.K., Dolan, M., Reeve, M.J., Rafalski, J.A., Rothman-Denes, L.B., & LaRossa, R.A. (2001). A genomic approach to gene fusion technology. Proc Natl Acad Sci 98, 2555–2560. Vasconcelos, G.J., & Swartz, R.G. (1976). Survival of bacteria in seawater using a diffusion chamber apparatus in situ. Appl Environ Microbiol 31, 913–920. Walkup, L.K., & Kogoma, T. (1989). Escherichia coli proteins inducible by oxidative stress mediated by the superoxide radical. J Bacteriol 171, 1476–1484. Wei, Y., Lee, J.M., Richmond, C., Blattner, F.R., Rafalski, J.A., & LaRossa, R.A. (2001). High-density microarraymediated gene expression profiling of Escherichia coli. J Bacteriol 183, 545–556. Welch, T.J., Farewell, A., Neidhardt, F.C., & Bartlett, D.H. (1993). Stress response of Escherichia coli to elevated hydrostatic pressure. J Bacteriol 175, 7170–7177. Willsky, G.R., Bennett, R.L., & Malamy, M. (1973). Inorganic phosphate transport in Escherichia coli: Involvement of two genes which play a role in alkaline phosphatase regulation. J Bacteriol 113, 529–539. Xiao, H., Kalman, M., Ikehara, K., Zemel, S., Glaser, G., & Cashel, M. (1991). Residual guanosine 3
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5 Human Pathogenic Viruses in the Marine Environment Albert Bosch, F. Xavier Abad, and Rosa M. Pint´o 5.1. INTRODUCTION Indigenous marine virus strains outnumber any form of life in the sea, usually occurring in billion amounts per liter (Danovaro et al., 2001; Fuhrman, 1999). However, although transspecific propagation of viruses may take place, the virus–host relationship tends to be quite constrained, and consequently human viruses are the only viral agents of public health concern in the marine environment. Pathogenic viruses are routinely introduced into marine and estuarine waters through the discharge of treated and untreated sewage, since current water treatment practices are unable to provide virus-free wastewater effluents (Rao & Melnick, 1986). It is estimated that the number of cases of gastrointestinal illness annually reported worldwide accounts for billions (Oh et al., 2003; Parashar et al., 1998). A good deal of these diarrheal cases are to some extent the result of the fecal contamination of the marine environment (Cabelli et al., 1982; Fattal & Shuval, 1989; Koopman et al., 1982; Moore et al., 1994), while enterically transmitted hepatitis outbreaks, i.e. hepatitis A and E, have been reported to be associated to water and shellfish (Bosch et al., 1991, 2001; Halliday et al., 1991; Melnick, 1957; Reid & Robinson, 1987). The maintenance and assessment of the virological quality and safety of marine water systems employed for recreating and seafood harvesting is of seminal importance in the prevention of diseases transmitted through the fecal–oral route, and may lead to significant reductions of economic losses due to the closures of tourists resorts and shellfish harvesting areas. For this purpose, it is imperative to trace and characterize the type and origin of fecal contaminants in order to assess the associated health threat and the required corrective measures.

5.2. VIRUSES AND DISEASES Present in sewage contaminated waters are well over 100 virus species able to cause a wide spectrum of illnesses in mankind including hepatitis, gastroenteritis, meningitis, fever, Albert Bosch, F. Xavier Abad, and Rosa M. Pint´o of Barcelona, Barcelona, Spain.

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Enteric Virus Laboratory, Department of Microbiology, University

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Table 5.1. Human enteric viruses with potential waterborne transmission. Genus

Popular name

Diseases caused

Enterovirus

Polio Coxsackie A, B

Paralysis, meningitis, fever Herpangina, meningitis, fever, respiratory disease, hand-foot-and-mouth disease, myocarditis, heart anomalies, rush, pleurodynia, diabetes? Meningitis, fever, respiratory disease, rush, gastroenteritis Hepatitis Unknown Gastroenteritis Gastroenteritis, respiratory disease, conjunctivitis Gastroenteritis Gastroenteritis Hepatitis Gastroenteritis Gastroenteritis Gastroenteritis, respiratory disease Gastroenteritis

Hepatovirus Reovirus Rotavirus Mastadenovirus Norovirus Sappovirus To be determined Mamastrovirus Parvovirus Coronavirus Torovirus

Echo Hepatitis A Human reovirus Human rotavirus Human adenovirus Norwalk-like virus Sapporo-like virus Hepatitis E Human astrovirus Human parvovirus Human coronavirus Human torovirus

rash, conjunctivitis, and may be diabetes or SARS (Table 5.1). However, few examples exist with epidemiological proof of waterborne transmission (Craun et al., 2002). In 1979, it was estimated that between 5 and 18 million people die every year from gastroenteritis, rotavirus being the most important viral agent transmitted through the fecal– oral route. In the developing countries the burden of rotavirus disease in children under 5 years of age has been estimated to be over 125 million cases annually, of which 18 million are severe cases, and 500,000–800,000 deaths in children under the age of 4 are attributable to rotavirus diarrhea (Oh et al., 2003; Parashar et al., 1998). In the developed world, mortality due to rotavirus infection is very low; however, it remains an important cause of morbidity and of hospitalization in young children. Additionally, hepatitis A accounts for around half the total number of hepatitis cases diagnosed worldwide, and some regions, such as some areas in the Mediterranean region, are still endemic for hepatitis A. Poliomyelitis, not too long ago the most feared viral disease, has not yet been eradicated, although this long pursued objective seems presently within reach. Norwalk-like viruses, the type species of genus Norovirus and formerly included in the SRSV (small round structured viruses), account for over 90% of foodborne gastroenteritis affecting children and adults (Lopman et al., 2003). Additionally, astroviruses were reported in 1996 to be second only to rotaviruses as a cause of hospitalization for childhood viral gastroenteritis (Glass et al., 1996), while adenoviruses and Sapporo-like viruses have also been recognized as significant etiological agents of epidemic nonbacterial diarrhea (Lopman et al., 2003). Hepatitis E virus, currently unclassified, is the primary cause in tropical and subtropical developing countries of an enterically transmitted non-A non-B hepatitis, with a mortality rate of up to 20% in pregnant women (Reyes, 1993; Schlauder & Mushahwar, 2001). In November 2002, the initial cases of the emerging disease denominated “severe acute respiratory syndrome” or SARS were reported (Ksiazek et al., 2003). Although the primary mode of transmission of the SARS coronavirus appears to be direct mucous membrane contact with infectious respiratory droplets and/or through exposure to fomites, at the time of writing

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Figure 5.1. Routes of transmission of enteric viruses to become contaminants of the marine environment.

this chapter the role of fecal–oral transmission is unknown. In spite that several coronaviruses are spread by this route, there is no current evidence that this mode of transmission plays a key role in the transmission of SARS, although there is a considerable shedding of the virus in stool, where it remains stable at room temperature for several days (Tsang, 2003).

5.3. THE FATE OF ENTERIC VIRUSES IN THE MARINE ENVIRONMENT The demands exerted by the expanding world population and industry make the marine environment increasingly susceptible to pollution from municipal sewage, industrial effluents, and agricultural wastes. Figure 5.1 illustrates the potential routes of transmission of enteric viruses to become contaminants of the marine environment. Seawater pollution control relies on secondary treatment of sewage and on the theoretically infinite dilution of wastes in the receiving waters. However, the marine environment, including oceans, has a finite ability to receive and recover from waste disposal practices, and certainly is incapable of unlimited waste assimilation. Viruses are shed in extremely high numbers in the feces of infected individuals: e.g., patients suffering from gastroenteritis or hepatitis may excrete from 105 to 1011 virus particles per gram of stool (Farthing, 1989). Viruses are present in high numbers in raw wastewater and current water treatment practices fail to ensure the complete removal of viral pathogens (Rao & Melnick, 1986 ); consequently, viruses become environmental pollutants. There are several routes by which viruses reach the sea, including direct discharge of treated or untreated

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Figure 5.2. Fate of enteric viruses in coastal marine environments. A. A heavily polluted river with abundant particulate material discharges into the sea, B. undisrupted marine sediment with the fluffy top layer where viruses accumulate, C. coquina clams and other bivalves readily adsorb pathogenic viruses within their edible tissues, and D. shellfish grown in areas receiving urban sewage contamination is responsible for outbreaks of gastroenteritis and infectious hepatitis.

sewage effluents, unintentional discharges by urban and rural run-off, waste input from boats, and via rivers when the discharges take place in fresh water. Mankind is exposed to enteric viruses in seawater mainly through the consumption of shellfish grown in contaminated waters, or to a lesser extent through recreational activities in sewage-polluted waters. The type of treatment will ultimately determine the concentration of pathogens in treated sewage and sludge, and their relative risk of disposal. An overview of the fate of enteric viruses in coastal environments is depicted in Figure 5.2. Domestic sewage may be disposed of directly in the marine environment by coastal outfalls or by dumping from barges, and may occur in the form of raw sewage, treated effluent, or sewage sludge. Virus concentrations of 5000– 100,000 PFU/l are commonly reported in raw sewage (Rao & Melnick, 1986), and may be greatly reduced during treatment; however an average of 50–100 PFU/l are normally found in effluents from water treatment plants (Rao & Melnick, 1986). In any case, viruses readily adsorb onto the abundant suspended solids present in sewage and are discharged solid-associated into marine environments (Fig. 5.2A). While viruses associated with small-size (6 µm) particles readily settle down in the bottom sediment. Viruses accumulate in the loose fluffy top layer of the compact bottom sediment (Fig. 5.2B) and are thereby protected from inactivation by natural or artificial processes (Rao et al., 1986a; Sobsey et al., 1988). Sediments in coastal seawaters act as reservoirs from which viruses may be subsequently resuspended by several natural or artificial phenomena. Shellfish (Fig. 5.2C), being filter feeders, tend to concentrate viruses and bacteria in their edible tissues, and concentrations of these microorganisms in

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shellfish may be much higher than those in the surrounding water. Shellfish grown and harvested from waters receiving urban contaminants (Fig. 5.2D) have been implicated in outbreaks of viral diseases, notably viral hepatitis and gastroenteritis (Bosch et al., 2001; Christensen et al., 1998; Halliday et al., 1991; Kingsley et al., 2002; Le Guyader et al., 1996).

5.3.1. Viruses in Seawater Viruses are potentially dangerous contaminants of the sea, as products of human activities. Table 5.2 lists some reports on the isolation of human enteric viruses in coastal and estuarine seawaters. Human pathogenic viruses are not only found in the millions of liters of variously treated human wastes dumped directly into coastal waters from sewage outfalls, but also from runoff from numerous storm drain sewers. Although the ocean offers relatively rapid dilution of treated sewage, it appears that dilution phenomena may not be enough to minimize the impact of virus pollution when contaminant affluents discharge at or within a short distance of the shoreline (Bosch et al., 1988a; Hugues et al., 1980). Based on this evidence, barges are employed to dump treated effluents at considerable distances from the shore. Nevertheless, sea bathing and water sport activities, which are becoming increasingly popular, may be nowadays not only expanded beyond the traditional summer seasons but also carried out at long distances from the shore. The full impact of viral pollution of recreational seawater on public health is hard to evaluate due to the technical and economical difficulties of virological studies. Nevertheless, epidemiology studies show significant risks associated with swimming near flowing storm drains. Specifically, the incidence of illnesses such as rashes, gastrointestinal disorders, and upper respiratory infections approximately doubled in those swimming at the mouth of a storm drain (Santa Monica Bay Restoration Project, 1996). Additionally, epidemiological data of waterborne illnesses also indicate that the common etiological agents are more likely to be viruses and parasitic protozoa than bacteria, and enteroviruses (polioviruses, coxsackie viruses, echoviruses), hepatitis A and E, adenoviruses, rotaviruses, and noroviruses have all

Table 5.2. Examples of reported isolations of human enteric viruses in seawater. Site

Virus type

Virus numbers/liter

Reference

Italy USA (Texas) USA (New York)

Enteroviruses Enteroviruses Poliovirus Echovirus Enteroviruses Adenoviruses Enteroviruses Enteroviruses Enteroviruses Enteroviruses Poliovirus Echovirus Rotaviruses

0.4 to 16 TCIDa50 0.01 to 0.44 PFUb 0 to 2.1 PFU

De Flora et al., 1975 Goyal et al., 1979 Vaughn et al., 1979

0.05 to 6.5 MPNCUc

Hugues et al., 1980

0.12 to 1.72 MPNCU 0.05 to 0.14 PFU 1 to 6 PFU 0.06 to 0.026 PFU 0.12 to 0.15 MPNCU

Finance et al., 1982 Schaiberger et al., 1982 Fattal et al., 1983 Rao et al., 1984 Lucena et al., 1985

0.007 to 2.6 PFU

Rao et al., 1986a

France Spain USA (Florida) Israel USA (Texas) Spain USA (Texas) a b c

TCID50 : Tissue culture infectious dose50 . PFU: Plaque forming units. MPNCU: Most probable number of cytopathic units.

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been associated with swimming-related illness (Cabelli et al., 1982; Fattal & Shuval, 1989; Koopman et al., 1982; Moore et al., 1994; Seyfried et al., 1985). In a routine study conducted on seawater, a widespread poliomyelitis vaccination program conducted in a coastal urban community was reflected in the levels of vaccinal poliovirus in seawater, which demonstrates that environmental monitoring of viruses provides useful evidence on the occurrence of infections, not only apparent but also unapparent, among the population (Lucena et al., 1986).

5.3.2. Viruses in Marine Sediment Table 5.3 shows some examples of the occurrence of human enteric viruses in marine sediments. Viruses accumulate in the fluffy sediment where the concentrations may be 10– 10,000 times greater than those found in water. Sediments represent a reservoir from which viruses may spread when the sediments are disturbed and resuspended in the water column. Resuspension commonly occurs as a result of natural different phenomena, among which are storms, dredging, and tides (Van Donsel & Geldreich, 1971). Resuspended solid-associated viruses can be transported from polluted to unpolluted recreational or shellfish growing waters and then pose health hazards (Rao et al., 1986a). Obviously, the resuspension of the microorganisms has a more dramatic effect on the quality of the overlaying water in shallow estuarine water than in deep coastal water, where the impact would not be significant because of the dilution in the large volume of overlaying water (Schaiberger et al., 1982). Viruses that may become associated with particulate material present in seawater survive for longer periods of time when adsorbed to sediments than when held in free suspension (De Flora et al., 1975; Rao et al., 1984, 1986a; Sobsey et al., 1988). Studies have indicated the presence of human enteric viruses in the sediments of polluted coastal waters (Bosch et al., 1988a; Bosch & Pint´o, 1992; De Flora et al., 1975; Rao et al., 1986a; Schaiberger et al., 1982). Table 5.3. Examples of reported isolations of human enteric viruses in marine sediment. Site

Virus type

Virus numbers/kilogram

Reference

Italy

Enteroviruses Reovirus Enteroviruses Enterovirus Rotavirus Enterovirus Enterovirus Rotavirus Rotavirus Hepatitis, A virus Enterovirus Rotavirus Hepatitis A virus

0.4 to 40 TCIDa50

De Flora et al., 1975

0 to 112 PFUb 39 to 398 PFU 800 to 3800 PFU 5 to 73 PFU 130 to 200 PFU 57 to 140 FFc 0 to 560 FF, + RNAd + RNA + RNA + RNA

Schaiberger et al., 1982 Rao et al., 1984 Rao et al., 1986a Bosch et al., 1988a Jofre et al., 1989

USA (Florida) USA (Texas) USA (Texas) Spain Spain Spain France

a

TCID50 : Tissue culture infectious dose50 . PFU: Plaque forming units. FF: Fluorescent foci. d RNA detected by molecular hybridization. b c

Bosch & Pint´o, 1992 Le Guyader et al., 1994

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Human pathogenic viruses have been detected as far as 5 km from the shore and as deep as 82 m (Bosch et al., 1988a). Usually, no correlation could be demonstrated between virus numbers and any bacterial parameter in marine sediment samples. This lack of correlation is probably due to the different decay rates shown by bacteria and viruses in the environment. Only in cases where the pollution may result from a recent deposition may viruses correlate with bacterial indicators (Bosch et al., 1988a). An evaluation of the presence of viruses in marine sediment certainly provides an additional insight into the long-term water quality conditions of a given area.

5.3.3. Virus Survival in the Sea It may be assumed that human and animal pathogens will suffer an aggressive treatment in the marine environment and will die-off within a short time. The phenomenon that selfpurification processes are more pronounced in seawater than in river water has been reported by several authors (Matossian & Garabedian, 1967; Giron´es et al., 1989). However, the selfdepuration capacity of marine water is finite. Yet the critical question is whether or not these viruses can survive long enough and in high enough concentration to cause disease in individuals who are in contact with polluted recreational water or who consume contaminated seafood. The survival of viruses in the marine environment is reviewed in Chapter 16. Among the many factors affecting virus survival in the marine environment are temperature (Bosch et al., 1993), virus association to solids (Gerba & Schaiberger, 1975; La Belle et al., 1980; Rao et al., 1984; Sobsey et al., 1988), and the presence of microbial flora (Gunderson et al., 1968; Fujioka et al., 1980; Toranzo et al. 1983; Giron´es et al., 1989). It seems reasonable to assume that environmental factors and the compositional make-up of seawater may be substantially different from one geographical location to another, which implies that different data of virus persistence are produced when the same viral strain is suspended in water sampled from different sites (Bosch et al., 1993). Several observations demonstrate the potential involvement of native marine microorganisms in the inactivation of viruses in marine habitats, although data on the successful isolation of microorganisms with virucidal properties are scarce (Fujioka et al., 1980; Girones et al., 1990; Bosch et al., 1993). Additionally, the ability of bacteria to inactivate viruses is usually lost while subculturing the microorganisms in the laboratory, although in a few cases these bacteria could be subcultured for over 1 year without losing their antiviral activity (Girones et al., 1990; Bosch et al., 1993).

5.4. VIRUSES IN SHELLFISH Several phenomena, such as flooding, treated and untreated polluted effluent discharges, or sewage runoff, can elevate microbial contaminants in shellfish habitats, and, since bivalves are filter feeders, these mollusks can become reservoirs of human pathogens. Other types of seafood such as crabs (Goyal, 1984) or shrimps (Botero et al., 1996) can accumulate viruses on their shells and carnivorous shellfish, such as lobsters or crabs, can feed on contaminated bivalves (Hejkal & Gerba, 1981), but their role in the transmission of viral diseases is unproven. Following the culinary tradition, bivalve shellfish is often consumed raw, like oysters and sometimes clams or cockles, or just lightly cooked, like most of other mollusks. This cooking

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Table 5.4. Examples of reported large (over 100 cases) outbreaks linked to shellfish consumption. Year

Country

Shellfish

No. of cases

1973 1976–1977 1978 1978 1980–1981 1982 1983 1983 1986

USA Great Britain Australia Australia Great Britain USA Great Britain Malaysia USA

1988 1999

Shanghai Spain

Oysters Clams Oysters Oysters Cockles Oysters Oysters Cockles Clams Oysters Clams Clams

265 800 2000 150 424 472 181 322 813 204 292 301 183

a b c

Responsible virus

Reference

HAVa SRSVb NVc NV NV NV SRSV HAV NV

Mackowiak et al., 1976 Appleton & Pereira, 1977 Murphy et al., 1979 Linco & Grohmann, 1980 O’Mahony et al., 1983 Richards, 1985 Gill et al., 1983 Goh et al., 1984 Morse et al., 1986

HAV HAV

Halliday et al., 1991 Bosch et al., 2001

HAV: Hepatitis A virus. SRSV: Small round structured viruses. NV: Norovirus.

habit, together with the fact that the whole animal including viscera is consumed, poses a major public health concern since shellfish act like passive carriers of human pathogenic viruses. Generally, commercial growth of shellfish species takes place in shallow, in-shore waters, which may receive occasional sewage pollution. The consumption of shellfish is very clearly linked to the transmission of enteric infections and epidemics have been recorded since medieval times in many countries (Lees, 2000). These outbreaks are normally caused by shellfish collected by unscrupulous professionals or careless private individuals, from areas where harvesting is prohibited. However, they also occur as a result of eating shellfish from authorized shellfishproducing areas, when these areas have been temporarily polluted (Mackowiak et al., 1976) and sanitary controls fail to provide a safe indication of viral pollution. The first reported association of viruses with shellfish-borne gastroenteritis infection was observed in the winter of 1976–1977 in the United Kingdom when cockles were epidemiologically linked to 33 incidents affecting nearly 800 people (Appleton & Pereira, 1977). SRSV particles, like those seen in outbreaks of winter vomiting disease, were observed by electron microscopy in a high proportion of patients feces. Regardless of the variety of healthsignificant viruses found in shellfish, norovirus and hepatitis A virus are the most relevant viral pathogens involved in shellfish-borne diseases (Table 5.4). Norovirus infections represent the vast majority of shellfish-related outbreaks, and hepatitis A is the most serious infectious disease caused by shellfish consumption. The United States Food and Drug Administration risk assessments estimate cases of norovirus gastroenteritis related to seafood consumption at some 100,000 per year (Williams & Zorn, 1997), and epidemics of hepatitis A caused by food occur 10 times more often than those caused by water, shellfish being the cause of more than 50% of reported cases (Cliver, 1985). Nevertheless, no shellfish-borne outbreak ever had the magnitude of the one reported in Shanghai in 1988, caused by hairy clams (Halliday et al., 1991). Bioaccumulation of viruses in the shellfish digestive tract is a very rapid phenomenon. Viruses are readily adsorbed to shellfish tissue within 1-h contact time and maximum virus levels may be observed after 6 h (Abad et al., 1997a). Adsorption of viruses on substrates such as feces, kaolinite, or unicellular algae considerably increases shellfish accumulation

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efficiency (Metcalf et al., 1979). Commercial heat treatment (cooking) is employed to reduce the levels of microbial contaminants in shellfish. Heat can render many viruses noninfectious; however the degree of cooking required to reliably inactivate viruses would probably render oysters unpalatable to consumers (McDonnell et al., 1997). Laboratory studies show that enteric viruses and notably hepatitis A virus may be found in cooked shellfish (Abad et al., 1997). In addition, outbreaks of gastroenteritis and hepatitis have been linked to consumption of commercially cooked cockles (Appleton & Pereira, 1977; Kohn et al., 1995). Most countries have endorsed sanitary controls on live bivalve shellfish. In the European Union, these are covered by Council Directive 91/492/EEC (Anonymous, 1991) and in the United States, by interstate trading agreements set out in the Federal Drug Administration National Shellfish Sanitation Program Manual of Operations (Anonymous, 1993). These regulations cover similar ground on the requirements, among others, for harvesting area classification, depuration, relaying, analytical methods, and provisions for suspension of harvesting from classified areas following a pollution or public health emergency. A major weakness of these controls is the use of traditional bacterial indicators of fecal contamination, such as the fecal coliforms or Escherichia coli, to assess contamination and hence implement the appropriate control measures. Fecal indicators are either measured in the shellfish themselves (EU perspective) or in the shellfish-growing waters (US FDA perspective). However, several reports describe a lack of correlation between bacterial indicator microorganisms and viruses, and pathogenic viruses may be detected in shellfish from areas classified as suitable for commercial exploitation according to fecal coliform criteria (Abad et al., 1997a; Lees, 2000; Le Guyader et al., 2000). The guidelines establish that shellfish meeting a microbiological standard of less than 230 E. coli or 300 fecal coliforms in 100 g of shellfish flesh can be placed on the market for human consumption. Human enteric viruses, e.g. rotavirus and hepatitis A virus, have been detected in shellfish which were adequate for public consumption according to criteria based on the numbers of bacterial indicators (Jofre et al., 1993; Bosch, et al., 1994; Le Guyader et al., 2000; Romalde et al., 2002). Additionally, hepatitis A and gastroenteritis outbreaks have been associated with the consumption of shellfish meeting legal standards (Bosch et al., 2001; Le Guyader et al., 1996; Lees, 2000; Mele et al., 1989). The legislation also requires that third country imports into the EU and USA have to be produced to the same standard as domestic products. Exporting nations have therefore developed programs for compliance with the regulations of their target export markets. Nevertheless, a number of examples of trans-national outbreaks have recently been reported following trade between EU Member States (Christensen et al., 1998) and importation of shellfish from third countries into the EU and the USA (Bosch et al., 2001; S´anchez et al., 2002; Kingsley et al., 2002). Conventional commercial processes employed to purge out the microbial contamination of live bivalves are depuration, performed in tanks, and relaying, performed in the natural environment. Tank-based depuration is now widely practiced in many European countries, while it is less widely used in the USA (Lees, 2000; Otwell et al., 1991). Depuration periods may vary from 1 to 7 days, since minimum time periods for depuration are not stipulated in the legislation, with around 2 days being probably the most widely used period. Early studies, using artificially spiked soft shell clams, reported that most viruses were purged within a 24–48-h period, and that low levels of viruses were depurated more rapidly than high levels (Metcalf et al., 1979). More recent studies show that although depuration

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and relaying procedures may be insufficient to completely remove viruses (Abad et al., 1997a; Lees, 2000; Richards, 1988), they do contribute to reduce virus levels and hence the risk of infection due to shellfish consumption (Bosch et al., 1994). Process temperature appears as an important factor for effective virus removal (Dore et al., 1998; Jaykus et al., 1994; Power & Collins, 1990), although a raise in depuration temperature may result in increased shellfish mortality. Nevertheless, epidemiological evidence reveals that enteric viruses can be transmitted to mankind after consuming shellfish which has been depurated (Gill et al., 1983; Heller et al., 1986; Sockett et al., 1993). Once again compliance with bacterial end-product standards does not provide a guarantee of virus absence, and bacterial depuration rates cannot accurately predict virus removal rates.

5.5. METHODS OF VIRUS DETECTION 5.5.1. Virus Concentration from Seawater and Marine Sediment The basic steps of the virological analysis of seawater are sampling, concentration, decontamination/removal of inhibitors, and specific virus detection. Concentration is a particularly critical step since the viruses may be present in such low numbers that concentration of the water samples is indispensable to reduce the volume to be assayed for viruses to a few milliliters or even microliters. A good concentration method should fulfill several criteria: it should be technically simple, fast, provide high virus recoveries, be adequate for a wide range of enteric viruses, provide a small volume of concentrate, and be inexpensive. Table 5.5 shows a broad selection of currently available and widely employed procedures; some of them require large equipment. Details on virus concentration procedures have been published elsewhere (American Public Health Association, 1998; Environmental Protection Agency, 1984). In relatively nonpolluted marine waters, the virus levels are likely to be so low that optimally hundreds, or even thousands, of liters should be sampled to increase the probability of virus detection. Nevertheless, all available methodologies have important limitations. The efficiency Table 5.5. Procedures for the concentration of viruses from seawater. Principle

Negatively charged filters Positively charged filters Glass powder Glass fiber Aluminum hydroxide

References Adsorption–elution methods Farrah et al., 1977 Sobsey & Jones, 1979 Gajardo et al., 1991; Sarrette et al., 1977; Schwartzbrod & Lucena, 1978 Vilagin`es et al., 1997 Wallis & Melnick, 1967

Precipitation methods Organic flocculation Katzenelson et al., 1976 Ammonium sulfate precipitation Bosch et al., 1988b; Shields & Farrah, 1986 Polyethylene glycol hydroextraction Farrah et al., 1978; Lewis & Metcalf, 1988 Ultracentrifugation Mehnert et al., 1997; Steinman, 1981 Lyophilization Gajardo et al., 1995; Pint´o et al., 2001 Ultrafiltration Divizia et al., 1989

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of a virus concentration method widely depends on the quality of the sampled water. Usually, all available procedures have been basically evaluated in spiked samples, and the recovering efficiencies recorded with experimentally contaminated water dramatically decrease when the method is applied in actual field trials. Additionally, none of the existing concentration procedures has been tested with all the groups of medically important viral pathogens; normally, a few specific enteric viruses have been employed to conduct the evaluation trials. However, some virus concentration methods have been used successfully to recover naturally occurring enteric viruses in seawater (Environmental Protection Agency, 1984; Finance et al., 1982; Gerba & Goyal, 1982; Goyal & Gerba, 1983; Henshilwood et al., 1998; Lewis & Metcalf, 1988; Rao et al., 1986a). Most of the procedures for concentrating and extracting viruses make use of the properties of the viral proteinaceous macromolecules. Certain protein structures confer on viruses in an aquatic environment the properties of a hydrophilic colloid of an amphoteric nature whose electric charge varies according to the pH and the ionic force of the environment. Viruses can therefore be adsorbed onto and then detach themselves from different substrates which are positively or negatively charged depending on their pH. Methods based on the adsorption of viruses from the sampled water onto a suitable solid surface from which they may be subsequently eluted into a much smaller volume are recommended for their use with largevolume samples. Different types of filters have been proposed for the recuperation of aquatic viruses, in the form of flat membranes or cartridges. Cartridge-type filters have the advantage of allowing filtration of large volumes of moderately turbid water within relatively short time. Their chemical composition, diameter, and porosity vary enormously. A whole range of “negatively” or “positively” charged filters now exist. Their efficiency depends on the type of water being treated and the presence of interfering substances such as detergents, suspended solid matter, or organic matter which can affect the adsorption of the viruses on these filters (Sobsey & Glass, 1984; Sobsey & Hickey, 1985). Negatively charged membranes or cartridges (Farrah et al., 1976; Wallis & Melnick, 1967) have the drawback that the water samples must be pretreated, i.e., the water has to be acidified and often salts need to be added to facilitate adsorption since electronegative filters do not adsorb viruses well under ambient water conditions (Rao & Melnick, 1986), and this limits the on-location use of this method to a great extent. However automatic injection systems do exist for treating several hundred liters of water. Virus concentration with electropositive filters may be performed on location, at ambient conditions, without any prior amendment of the sample, which make this procedure most suited for in-field studies, provided that the sample pH is lower than 8.5 (Sobsey & Jones, 1979). Glass powder (Gajardo et al., 1991; Sarrette et al., 1977; Schwartzbrod & Lucena, 1978) or glass fiber (Vilagin`es et al., 1997) has also been satisfactorily used in different laboratories as adsorbent materials for virus concentration. Viruses in eluate volumes too large to be conveniently and economically assayed directly for specific virus detection, such as those obtained from processing large volumes of water through cartridge or large disk filters, can be reconcentrated by several alternative methods. Recovery of small quantities of viruses from natural waters is dependent not only on the efficacy of primary concentration from the original large volume, but also on the reconcentration of the primary eluate to a smaller volume. Methods such as aluminum hydroxide adsorption–precipitation (Wallis & Melnick, 1967), polyethylene glycol hydroextraction (Farrah et al., 1978; Lewis & Metcalf, 1988), organic

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flocculation (Katzenelson et al., 1976), and ammonium sulfate precipitation (Bosch et al., 1988; Shields & Farrah, 1986) that are impractical for processing large fluid volumes are however suitable for second-step concentration. Alternatively for reconcentration, viruses can also be sedimented depending on their molecular weight, and certain methods of concentration, such as ultracentrifugation (Mehnert et al., 1997; Steinman, 1981), take advantage of this characteristic. Freeze-drying of samples (Gajardo et al., 1995; Pint´o et al., 2001) and rehydration in a smaller volume provides a procedure for both virus concentration and removal of inhibitors, if molecular procedures are employed for virus detection. Ultrafiltration (Divizia et al., 1989) can utilize size exclusion rather than adsorption and (or) elution to concentrate viruses and, therefore, can provide consistent recoveries among different viruses and widely varying water conditions. Since an evaluation of the presence of viruses in sediment provides an additional insight into long-term water quality conditions, several methods for the detection of viruses have been developed. These methods basically consist of an elution of the viruses from the solid materials and an ulterior concentration of the eluted viruses. Viruses are usually eluted from the marine sediments by using alkaline buffers (Bosch et al., 1988a; Gerba et al., 1977; Jofre et al., 1989), or by the action of caotropic agents (Jofre et al., 1989; Lewis & Metcalf, 1988; Wait & Sobsey, 1983). Precipitation methods based on organic flocculation (Wait & Sobsey, 1983), ammonium sulfate flocculation (Jofre et al., 1989), polyethylene precipitation (Lewis & Metcalf, 1988), or ultrafiltration (Gerba et al., 1977) are procedures commonly employed to concentrate viruses from the eluate.

5.5.2. Virus Recovery from Shellfish Virus detection in shellfish has to overcome several difficulties. On the one hand, viruses are expected to be present in shellfish in very low numbers, which nevertheless are sufficient to pose a health risk. This low virus load implies the use of methodologies yielding a high efficiency of virus recovery from shellfish tissues. On the other hand, shellfish extracts are both highly cytotoxic and not adequate to be inoculated in cell cultures for the detection of culturable viruses, and not compatible either with polymerase chain reaction (PCR) based methodologies for the detection of nonculturable viruses, particularly if a reverse transcription must be previously performed (RT-PCR). The key objective is then to develop procedures for shellfish analysis which result in a low volume of a noncytotoxic or, even better nowadays, highly pure nucleic acid preparation with no inhibitory effect to the PCR. As a matter of fact, in this latter case, the degree of virus detection effectiveness achieved after RT-PCR is essentially the result of two related factors: the efficiency of recovery of the extraction procedure applied to the shellfish sample and the degree of final purity of the recovered virus. Table 5.6 lists different procedures for the processing of shellfish samples prior to the specific virus detection, essentially by molecular procedures since the most relevant shellfishborne viral pathogens are nonculturable. The first decision is to choose between performing virus detection in dissected shellfish tissues or in whole shellfish meats. Studies on the localization of human enteric viruses in shellfish tissues revealed that most of the virus could be found in the stomach and digestive diverticula (Abad et al., 1997a; Romalde et al., 1994). Atmar and co-workers reasoned that removal of these organs for virus extraction might simplify and shorten the time needed to purify viral nucleic acid for RT-PCR (Atmar et al., 1996). Testing the stomach and digestive gland for virus detection presented several advantages in comparison

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Table 5.6. Procedures for processing of shellfish samples prior to virus detection by RT-PCR. Process

Method

Reference

Virus elution

Chloroform-butanol/cat-floc elution Vertrel extraction Organic flocculation Centrifugation Ultracentrifugation PEG precipitation Guanidium thiocyanate Cetyltrimethyl ammonium bromide Commercial nucleic acid extraction kits

Atmar et al., 1996 Mendez et al., 2000 Sobsey et al., 1978 Sobsey et al., 1978 Loisy et al., 2000 Lewis & Metcalf, 1988; Atmar et al., 1995 Boom et al., 1990; Lees et al., 1994 Atmar et al., 1995; Jaykus et al., 1994 Loisy et al., 2000; Schwab et al., 2000; Shieh et al., 1999

Virus concentration

RNA extraction

with testing whole shellfish: less time-consuming procedure, increased test sensitivity, and decrease in the sample-associated interference with RT-PCR. Following virus extraction, a variety of subsequent nucleic acid extraction and purification protocols may be employed (Table 5.6). Due to the small size of the PCR reaction volumes, a reconcentration step is incorporated prior to the molecular assay. Nucleic acid purification based on virus lysis with guanidine and recovery with a silica matrix (Boom et al., 1990; Lees et al., 1994), or, alternatively, the use of organic solvents for purification, followed by nucleic acid precipitation using cetyltrimethyl ammonium bromide (CTAB) (Atmar et al., 1995; Jaykus et al., 1994), remain the procedures of choice. However, a wide variety of commercial kits have been applied for nucleic acid purification, offering reliability combined with convenience (Loisy et al., 2000; Schwab et al., 2000; Shieh et al., 1999).

5.5.3. Specific Virus Detection Procedures Not too long ago, methods for the detection of viral pathogens were restricted to assays for culturable viruses, focused almost entirely on enteroviruses, and the BGM cell line has been for long the choice for infectivity assays of enteroviruses in environmental samples (Bosch, 1998; Morris & Waite, 1980; Rao et al., 1986b). Despite that enteroviruses do not appear as epidemiologically relevant environmental contaminants; it will remain important to gather data on their occurrence in the environment until the global eradication of poliomyelitis becomes a reality. However, for this latter purpose, molecular tools provide better perspectives than cell culture techniques. Wild-type rotaviruses present difficulties in their in vitro replication, although most isolates may be adapted to grow in several cell lines such as the monkey kidney cell line MA104 or the human intestinal cell line CaCo-2 (Kitamoto et al., 1991). The standard methods for the diagnosis of specific infectious rotaviruses involve immunofluorescence tests and optical microscopic counting of infected foci in the culture (Bosch et al., 1988b; Hejkal et al., 1984; Smith & Gerba, 1982). A further refinement in this direction was the use of flow cytometry for the detection of fluorescent foci in rotavirus infected cells (Abad et al., 1998). Flow cytometry is applicable for the detection of rotaviruses in environmental samples through an automatable and standardizable procedure, much less cumbersome than direct optical microscopy screening of cell cultures for fluorescent foci.

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The advent of molecular techniques for virus detection, and particularly reverse transcriptase-polymerase chain reaction (RT-PCR), provided exquisite tools for the detection of fastidious health-significant viruses in the water environment. Health-significant viruses, which were previously unrecognizable because they replicate poorly or do not at all in cell cultures, became detectable with nucleic acid based techniques. Environmental virologists initially employed hybridization assays which have been more recently replaced by PCRbased procedures (Bosch et al., 1996; Jaykus et al., 1996; Jothikumar et al., 1995; Le Guyader et al., 1994; Lees et al., 1994; Pint´o et al., 1996; Schwab et al., 1995; Villena et al., 2003). Enhanced detection sensitivity, a significant advantage in environmental virology, may be achieved through nested approaches (Dore et al., 1998; Green et al., 1998; Le Guyader et al., 1994). However, the fact that nested PCR may represent a frequent cause of carryover contamination should not be neglected. An additional difficulty is that traditional virus concentration systems from water or shellfish extraction procedures are not always compatible with RT-PCR detection: inhibitory substances are concentrated and recovered along with the viruses. A great variety of procedures have been developed for the removal of inhibitors, which include dialysis, solvent extraction, proteinase treatments, lyophilization, gel or glass filtration, nucleic acid adsorption or precipitation, antibody capture, and the use of commercial kits (Abbaszadegan et al., 1993; Atmar et al., 1995; Graff et al., 1993; Jaykus et al., 1996; Loisy, et al., 2000; Schwab et al., 2000; Shieh et al., 1999; Tsai et al., 1993). The rule of thumb is that the degree of final purity of the assayed sample greatly determines the sensitivity of PCR, or particularly, RT-PCR virus detection. One limitation of molecular techniques is that they fail to discern between infectious and noninfectious particles which may be of critical relevance in environmental virology (Abad et al., 1994; Gassilloud et al., 2003). All enteric viruses of public health concern bear RNA genomes. In studies employing RT-PCR, it has been shown that poliovirus genomic RNA is not stable in nonsterilized seawater (Tsai et al., 1995). While free DNA is fairly stable, it is unlikely that a free single-stranded RNA genome like those of noroviruses or hepatitis A virus would remain stable without its protein coat in the marine environment. This presumption is less clear for the double-stranded RNA genome of rotaviruses. One possibility of solving this problem may be the use of cell lines susceptible to support the propagation of a wide variety of enteric viruses, enabling the amplification of virus sequences in cell culture prior to detection by PCR, accomplishing the dual purpose of increasing the number of copies of target nucleic acid and of incorporating an infectivity assay as well (Ma et al., 1994; Pint´o et al., 1995). This approach has been reported for detection of infectious astrovirus (Abad et al., 1997b) and enterovirus (Murrin & Slade, 1997; Reynolds et al., 1996). Whenever possible, the use of a combined cell culture-RT-PCR procedure utilizes the major advantages of the separate methodologies, while overcoming many of their disadvantages. The inclusion of an infectivity test prior to the specific detection solves not only this latter point but also the lack of sensitivity required for some types of samples such as environmental samples. However, combined cell culture-RT-PCR methods are not applicable to the unculturable caliciviruses or to most hepatitis A virus isolates. An alternative approach is to employ an antibody capture RT-PCR. This has been applied to detection of hepatitis A virus in seeded shellfish samples and shown to be both sensitive and useful for removal of RT-PCR inhibitors as well (Deng et al., 1994; Graff et al., 1993; L´opez-Sabater et al., 1997). Since recognition by

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a conformationally dependent monoclonal antibody is lost when the particle conformation is altered, coupling of the molecular procedure with capture with this type of antibody may enable to discern between intact and altered virions. This approach may prove useful for other enteric viruses, provided that adequate immunological reagents for the most relevant viral pathogens are available. For this purpose, recombinant virus-like particles, which can be obtained in very high numbers in in vitro expression systems (Crawford et al., 1994; Lawton et al., 1997), may be employed for the production of antibodies of non culturable viruses. Many potential users may find PCR cumbersome, since a single test entails many different manual steps, and will consider the technique as suitable only for academic or reference labs, and inadequate for routine monitoring. However, over the last decade, PCR technology improved on several fronts. On the one hand, commercial PCR systems significantly ameliorated convenience, and have been quickly adopted for diagnostic laboratories. Nevertheless, the most dramatic improvement comes from the emergence of combined rapid thermocycling and fluorescence monitoring of amplified product, collectively referred to as “rapid-cycling real-time PCR” (Cockerill, III & Smith, 2002; Gardner et al., 2003; Gassilloud et al., 2003; Kageyama et al., 2003), together with nucleic-acid-sequence-based amplification or NASBA techniques (Jean et al., 2001; Yates et al., 2001), both of which are now applicable in several commercially available systems. These procedures enable not only qualitative determination but also, and particularly, quantitative diagnostic assays. Although the generic determination of pathogens is the essence of diagnostic practices, the possibility to quantitatively detect virus agents represents a seminal refinement in routine monitoring virology.

5.6. THE PROSPECTS FOR VIRUS STUDIES IN SEAWATER AND SHELLFISH The last two decades of virological research have contributed to significant progresses in the field of medical virology. These include the development of methodologies for the detection and characterization of nonculturable waterborne and foodborne viruses, the recognition of waterborne outbreaks caused by hepatitis A and E viruses, the consideration of rotavirus as the single most important cause of severe children gastroenteritis and norovirus as the most frequent agent of foodborne diarrhea, the characterization of other important agents of nonbacterial gastroenteritis such as astroviruses, sappoviruses, adenoviruses, and the assessment of the zoonotical transmission of some of the aforementioned agents. Molecular characterization of environmental virus isolates provides tools to acquire an overview of the epidemiology of viruses circulating in the community and, at the same time, unveil the occurrence of asymptomatic infections. Limitations of this approach are important environmental virology issues such as the differential stability of a given virus type could give rise to a certain level of distortion in the produced data, and the kind of sample analyzed: only those viruses that are more prevalent in the population, and thus excreted in higher numbers, are likely to be detected. Nucleic-acid-amplification-based techniques have been, and undoubtedly will continue to be, a major step forward in virus monitoring in environmental samples, and specially in shellfish and shellfish waters. One major drawback is that current published methods are diverse, complex, poorly standardized, and restricted to a few specialist laboratories. Standardized PCR assays for detection of public health relevant viruses as hepatitis A virus and

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norovirus are needed to be used in a routine diagnostic context. However, for this purpose, molecular techniques for the virological analysis must put in practice unique or additional quality assurance and quality control requirements to ensure that the data generated are useful and reliable. Some laboratories already have their own internal criteria for quality assurance but a consistent set of quality assurance procedures should be implemented in order to generate data that could be comparable and reliable. A major obstacle for acceptance of molecular techniques by regulatory agencies is in particular a lack of confidence on data generated without adequate controls. Availability of quantitative and standardized virus methods will enable the future setting of legislative virus standards for bathing waters, bivalve shellfish, and shellfish growing waters. Achievement of this objective will also enable the identification of key environmental factors, such as rainfall and sewage discharges, responsible for viral contamination in bathing and shellfish harvesting areas. Identification, and management, of such critical control points will provide an alternative approach to containing the virus risk and would permit the development of enhanced sanitary controls. In the meantime and due to technical difficulties, tests for most of these viruses remain restricted to laboratories with sophisticated facilities and well-trained personnel. On the other hand, it is impracticable to monitor the presence of all viral pathogens in the environment. The unreliability of bacterial model mcroorganisms led to the search for alternatives, and several bacteriophage groups appear as promising candidates, among which are somatic coliphages (IAWPRC Study Group on Health Related Water Microbiology, 1991), F+ specific (male-specific), RNA bacteriophages (Havelaar, 1993), and Bacteroides fragilis bacteriophages (Tartera & Jofre, 1987), all of them with available ISO (International Standardization Office) procedures for their detection in water. F+ phages in particular have been described as promising candidates to evaluate the virological quality of shellfish (Chung et al., 1998; Lees, 2000). Several studies have shown a correlation between the elimination kinetics of F+ RNA phages and those of enteric viruses (Dore & Lees, 1995; Power & Collins, 1989, 1990). F+ RNA phages were considered for inclusion in EU legal standards for live bivalve mollusk purification. Nevertheless reports on discrepancies in the occurrences of F+ phages and pathogenic viruses are frequent. In shellfish associated with a large outbreak of hepatitis A reported in the East of Spain in 1999, with 184 serologically confirmed cases, the discrepancy observed between hepatitis A virus and F+ phages was 55%, while a 50% discordance was ascertained between generic enteric virus occurrence and F+ presence (Bosch et al., 2003). In another study comparing the validity of E. coli, enterovirus, and F+ RNA bacteriophages as indicator microorganisms, the phages failed to predict the risk of viral illness (Miossec et al., 2001). Additionally, when the comparative positivity for human enteric viruses and F+ RNA phages was investigated in 101 randomly chosen shellfish samples from South and West coast of France, a good correlation between the occurrence of enteric viruses and F+ phages was observed in only 49% of the samples (Le Guyader, unpublished results). Forty one percent of these samples were positive for at least one type of enteric virus but negative for F+ phages, while 11% of the samples were found to be negative for enteric viruses but positive for F+ phages. At the time of publication of this work, F+ RNA phages have been removed from EU standards for shellfish purification. Exhaustive studies are still required to ascertain the validity of a candidate indicator in a given scenario. In the end we should probably give up our hopes of finding a “universal” indicator for viruses, applicable to all situations, and resign ourselves to the use of particular indicator, index, or model microorganisms for specific purposes.

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Finally, another long-time challenge in environmental virology is to conduct actual field studies to evaluate the environmental behavior of human enteric viruses, which has to face the impossibility of introducing pathogens in the environment. As model systems, recombinant tracers could be perfectly adequate for field studies of microbial tracking, since they may be produced in extremely high numbers (several milligram amounts). Additionally, their noninfectious nature makes them completely harmless and suitable to be used in scenarios where the use of actual viruses is hampered by the impossibility of introducing potential pathogens into, for instance, shellfish growing waters. Recombinant virus-like particles have been employed to investigate the influence of electrostatic interactions in the filtration of norovirus in quartz sand (Redman et al., 1997) and the behavior of rotaviruses in disinfection studies (Caballero et al., 2004). Obviously, from the strictly structural point of view, there is no better surrogate of an actual virus pathogen to track their behavior in the environment than a noninfectious virus-like particle of the same virus.

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Seyfried, D.L., Tobin, R.S., Brown, W.E. & Ness, P.F. (1985). A prospective study of swimming-related illness. II. Morbidity and the microbiological quality of water. Am J Public Health 75, 1071–1075. Shieh, C.Y.-S. Calci, K.R., & Baric, R.S. (1999). A Method To Detect Low Levels of Enteric Viruses in Contaminated Oysters. Appl Environ Microbiol 65, 4709–4714. Shields, P.A., & Farrah, S.R. (1986). Concentration of viruses in beef extract by flocculation with ammonium sulphate. Appl Environ Microbiol 51, 211–213. Smith, E.M., & Gerba, C.P. (1982). Development of a method for detection of human rotavirus in water and sewage. Appl Environ Microbiol 43, 1440–1450. Sobsey, M.D., Carrick, R.J., & Jensen, H.R. (1978). Improved methods for detecting enteric viruses in oysters. Appl Environ Microbiol 36, 121–128. Sobsey, M.D. & Glass, J.S. (1984). Influence of water quality on enteric virus concentration by microporous filter methods. Appl Environ Microbiol 47, 956–960. Sobsey, M.D., & Hickey, A.R. (1985). Effect of humic and fulvic acid on poliovirus concentration from water by microporous filtration. Appl Environ Microbiol 49, 259–264. Sobsey, M.D., & Jones, B.L. (1979). Concentration of poliovirus from tap water using positively charged microporous filters. Appl Environ Microbiol 37, 588–595. Sobsey, M.D., Shields, P.A., Hauchman, F.S., Davis, A.L., Rullman, V.A. & Bosch, A. (1988). Survival and persistence of hepatitis A virus in environmental samples. In A.J. Zuckerman (ed), Viral Hepatitis And Liver Disease Alan R. Liss, New York, pp. 121–124. Sockett, P.N., Cowden, J.M., LeBaigue, S., Ross, D., Adak, G., & Evans, H. (1993). Foodborne disease surveillance in England and Wales: 1989–1991. Commun Dis Rep 3, 159–174. Steinman, J. (1981). Detection of rotavirus in sewage. Appl Environ Microbiol 41, 1043–1045. Tartera, C., & Jofre, J. (1987). Bacteriophages active against Bacteroides fragilis in sewage-polluted waters. Appl Environ Microbiol 53, 1632–1637. Toranzo, A.E., Barja, J.L., & Hetrick, F.M. (1983). Mechanism of poliovirus inactivation by cell-free filtrates of marine bacteria. Can J Microbiol 29, 1481–1486. Tsai, Y.L., Sobsey, M.D., Sangermano, L.R., & Palmer, C.J. (1993). Simple method of concentrating enteroviruses and hepatitis A virus from sewage and ocean water for rapid detection by reverse transcriptase-polymerase chain reaction. Appl Environ Microbiol 59, 3488–3491. Tsai, Y.-L., Tran, B., & Palmer, C.J. (1995). Analysis of viral RNA persistence in seawater by reverse transcriptasePCR. Appl Environ Microbiol 61, 363–366. Tsang, T. (2003). Environmental issues. Presented at WHO Global Conference on Severe Acute Respiratory Syndrome (SARS), Kuala Lumpur, Malaysia, June 17–18 2003. Van Donsel, D.J., & Geldreich, E.E. (1971). Relationships of Salmonellae to fecal coliforms in bottom sediments. Water Res 5, 1079–1087. Vaughn, J.M., Landry, E.F., Thomas, M.Z., Vicale, T.J., & Penello, W.F. (1979). Survey of human enterovirus occurrence in fresh and marine surface waters on Long-Island. Appl Environ Microbiol 38, 290–296. Vilagin`es, P., Sarrette, B., Champsaur, H., Hugues, B., Dubrou, S., Joret, J.-C., Laveran, H., Lesne, J., Paquin, J.L., Delattre J.M., et al. (1997). Round robin investigation of glass wool method for poliovirus recovery from drinking water and sea water. Water Sci Technol 35, 445–449. Villena, C., Morsy El-Senousy, W., Abad, F.X., Pint´o, R.M., & Bosch, A. (2003). Group A rotavirus in sewage samples from Barcelona and Cairo: Emergence of unusual genotypes. Appl Environ Microbiol 69, 3919–3923. Wait, D.A., & Sobsey, M.D. (1983). Method for recovery of enteric viruses from estuarine sediments with caotropic agents. Appl Environ Microbiol 46, 379–385. Wallis, C., & Melnick, J.L. (1967). Concentration of viruses on aluminium and calcium salts. Am J Epidemiol 85, 459. Williams, R.A., & Zorn, D.J. (1997). Hazard analysis and critical control point systems applied to public health risks: The example of seafood. Revue Scientifique et Technique, 16, 349–358. Yates, S., Penning, M., Goudsmit, J., Frantzen, I., Van der Weijer, B., Van Strijp, D., & Van Gemen, B. (2001). Quantitative detection of hepatitis B virus DNA by real-time nucleic acid sequence-based amplification with molecular beacon detection. J Clin Microbiol 39, 3656–3665.

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6 Survival of Viruses in the Marine Environment Charles P. Gerba 6.1. INTRODUCTION It has been well established that human pathogenic viruses may be transmitted through the marine environment due to the release of sewage by polluted rivers, outfalls, or release from vessels. These wastes contain human enteric viruses, which if ingested, or in some cases inhaled, can cause a wide variety of illnesses. Their ability to be transmitted by this route is because of their capability to remain infectious long enough in the marine environment to come in contact with a susceptible host. Transmission routes may be fairly direct, such as ingested of contaminated seawater by a swimmer, or more complex by prolonged survival in sediments which are later resuspended and accumulated in shellfish during feeding. The virus is then transmitted during consumption of the shellfish. To understand the potential for human enteric virus transmission through the marine environment numerous studies have been conducted on factors which influence their persistence in this environment (Table 6.1). This review focuses on factors that could play a role in the survival of human pathogenic viruses in the marine environment.

6.2. FACTORS AFFECTING VIRUS SURVIVAL IN MARINE WATER 6.2.1. Types of Viruses All of the human enteric viruses transmitted by water and food are composed only of a nucleic acid genome of either RNA or DNA enclosed in a protective protein coat. There are currently over 140 different human enteric viruses that are known. Most of these have a genome of a single strand of RNA (sRNA), e.g. enteroviruses, astroviruses, caliciviruses, hepatitis A, hepatitis E. Rotaviruses have a double strand of RNA (dsRNA), and adenoviruses a double strain of DNA (dsDNA). The rotaviruses and adenoviruses are about twice the size (60 to 70 nm) of the sRNA viruses. Coronaviruses, picobirnavirues, and parvoviruses have been associated with gastroenteritis transmission but transmission by water or food has not been clearly demonstrated. Coronavirus (sRNA) are the only human enteric viruses with a lipid coat. The recently discovered picobirnavirues are small double stranded RNA viruses associated with Charles P. Gerba

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Department of Soil, Water and Environmental Science, University of Arizona, Tucson, AZ, USA.

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Table 6.1. Factors affecting virus survival in natural waters. Factor

Effect

Remarks

Temperature

Protein or nucleic acid denaturation

Light

Configuration change or breakage of the nucleic acid

pH

Release of nucleic acid or disassembly of capsid Stabilization of capsid structure

Temperature is probably the most important factor in virus survival; viruses survive longer at lower temperatures UV in sunlight and natural photodynamic processes may act upon the virus in natural waters Most enteric viruses are stable at pH values found in most natural waters Some viruses are protected against heat inactivation by the presence of certain cations; the reverse is also true Viruses survive longer in the presence of sewage and other soluble organic material Virus survival is prolonged by adsorption to clays and other suspended solids, viruses may also be inactivated by adsorption to other solid surfaces Certain viruses are inactivated more readily in the presence of aquatic bacteria; can also act to protect virus by adsorption to surface or other factors

Salts

Proteinaceous material Suspended particulate matter; sediments

Biological factors

Protection from enzymes and sunlight? Thermostabilization of protein capsid; protection from the action of enzymes Enzymes; adsorption; nonenzymatic interaction affecting capsid structure

diarrhea in immunocomprimised persons and travelers to developing countries. Parvoviruses are small (18 to 25 nm) singled stranded DNA viruses. They are among the most thermally resistant viruses known. Most of our understanding on virus survival in the marine environment before 1990 comes from studies with enteroviruses (sRNA) and bacteriophages of Escherichia coli (coliphages), although hepatitis A and noroviruses (a calicivirus) have been the viruses usually associated with transmission through the marine environment. This is because noroviruses have not yet been grown in the laboratory. During the last decade a great deal of information has been gained about the ecology and survival of phages of marine microorganisms (Wommack & Colwell, 2000). This has provided new information on potential mechanisms of virus persistence that may be useful in understanding the survival of human pathogenic viruses. However, it must be remembered that human enteric viruses are not native to this environment and the relative importance of factors affecting inactivation may be different between the two groups of viruses.

6.2.2. Temperature While many factors are involved in the survival of viruses in natural waters, temperature probably plays the most decisive role. Temperature is the only well-defined factor with a consistent effect on virus survival. At freezing or near-freezing temperatures, viruses may remain infectious for many months (Lo et al., 1976). Compared to many other viral groups that infect man, enteric viruses are usually heat stable. Hepatitis A virus, adenoviruses and parvoviruses appear to be the most thermostable enteric viruses (Crance et al., 1998; Brauniger et al., 2000; Gerba, unpublished observations).

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In addition, viral populations are heterogeneous and include thermo-resistant mutants (Pohjanpelto, 1961). The effect of temperature on the rate of viral inactivation is also related to factors such as pH, salts, organics, and the presence of particulate matter (Pohjanpelto, 1961; Salo & Cliver, 1976; Wallis et al., 1965; Wallis & Melnick, 1962). Certain cations in solution appear to protect some viruses from thermoinactivation, while having the opposite effect on other viruses. Thus, rotavirus SA-11 is more rapidly inactivated by heating at 50◦ C in 2 M MgCl2 , while reovirus type 1 infectivity is stabilized (Estes et al., 1979). In contrast, SA-11 infectivity is stabilized at 50◦ C in the presence of 2 M MgSO4 . At high temperatures, poliovirus type 1 is more stable in seawater than in distilled water (Liew & Gerba, 1980), indicating that salts in natural waters may have some stabilizing effect on virus survival. Poliovirus type 1 has been shown to be thermostablized by association with sediments (Liew & Gerba, 1980). At 50◦ C, poliovirus type 1 in artificial seawater is protected from inactivation in the presence of marine sediment. Almost all of the virus remained adsorbed to the sediment. The virus was also protected against inactivation at 24 and 37◦ C. Virus inactivation at elevated temperatures may result primarily from protein denaturation. Poliovirus RNA becomes more susceptible to ribionuclease after it is heated at high temperatures, indicating structural alterations in the viral capsid (Dimmock, 1967). At lower temperatures, inactivation due to damage of the RNA may also be important.

6.2.3. pH At the pH of most natural waters (pH 5 to 9), enteric viruses are very stable. Most enteric viruses are generally more stable at a pH between 3 to 5 than at alkaline pH (9 to 12). Adenoviruses and rotaviruses are readily inactivated at pH 10, while enteroviruses can survive at a pH of 11 to 11.5 for short periods of time (Gerba & Goyal, 1982). The pH may also greatly affect the mode of viral inactivation at different temperatures (Salo & Cliver, 1976).

6.2.4. Light While wavelengths of ultraviolet light in sunlight are known for their ability to inactivate viruses by causing cross-linking among the nucleotides, earlier studies did not suggest that they played a significant role in virus inactivation (Kapuscinski & Mitchell, 1980; Wommack et al., 1996). However, more recent studies on phage indigenous (virioplankton) to the marine environment suggests that they could play a significant role even at depths as great as 200 m (Suttle & Chen, 1992; Wommack & Colwell, 2000). Even in turbid estuarine waters, a modest effect of sunlight on viral decay can occur at a depth of 2.5 m (Cottrell & Suttle, 1995). Light wavelengths of 10 mg/kg; Aune et al., 2002). Arguments have thus been made for the use of analytical methods (rather than in vivo bioassays) to distinguish between these two toxin classes for regulatory purposes. The primary consequence of the more toxic i.p. yessotoxin exposure in mice (ca. 100 µg/kg) is reported as cardiotoxicity. Investigation of the yessotoxin mode of action is ongoing, with recent studies revealing a calcium-dependent activation of phosphodiesterase (Alfonso et al., 2003).

8.3.5. Dinoflagellates (Dinophysis), Pectenotoxins In addition to the production of DSP toxins (i.e., okadaic acid and dinophysistoxin) (see page 173), some of the same species of Dinophysis (e.g., D. fortii, D. acuta; see Landsberg, 2002) have also been implicated as producers of PTXs. Pectenotoxins are a group of at least ten cyclic polyether macrolides that also include recently identified “seco acid” congeners, which are produced enzymatically in shellfish (Quilliam, 2003a). Note that findings linking pectenotoxins production with Dinophysis species (as with production of dinophysistoxins) are based necessarily on natural samples, since this dinoflagellate genus remains recalcitrant to attempts at isolation into laboratory culture. Similar to the yessotoxins, pectenotoxins are also extracted together with dinophysistoxins and hence their association with this group of lipophilic toxins. In contrast to the dinophysistoxins, the pectenotoxins are strongly hepatotoxic (250 µg/kg i.p. in mice), although there is one report of diarrheagenic effects elicited by PTX-2 following oral exposure in mice, and reported to be comparable in toxicity to i.p. injection (Fern´andez et al., 2003). Data on the mechanism of action are scarce; however, recent work by Leira et al. (2002) on PTX-6 (a derivative of PTX-2) suggests that cytoskeletal disruption through depolymerization of F-actin may play an important role in toxicity.

8.3.6. Cyanobacteria (Lyngbya and Nodularia), Multiple Toxins, Human Health risks, Animal Mortalities Cyanobacteria blooms in freshwater systems cause significant public and animal health problems when organisms are exposed to cyanotoxins through drinking water from contaminated sources. In marine and brackish water systems, cyanobacteria such as Nodularia, Lyngbya, and Trichodesmium are less problematic, but nonetheless can cause localized problems (Sellner, 1997). There have probably been more toxic compounds identified from Lyngbya majuscula than from any other microalgal species (Landsberg 2002; Osborne et al., 2001). Most of the toxic compounds have been verified through experimental study, and it is unclear to what extent these compounds may affect aquatic organisms in the natural environment. It has been

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speculated that many of the bioactive metabolites produced by Lyngbya protect this microalga from predation by a variety of groups (e.g., crustaceans, herbivorous fish, and gastropods) (Orjala et al., 1995). L. majuscula is common in marine tropical areas and has principally been associated with contact skin dermatitis, eye, and respiratory irritation in humans swimming, bathing, or walking near affected waters (Grauer & Arnold, 1961; Osborne et al., 2001). Lyngbya has caused numerous outbreaks of dermatitis in the Hawaiian Islands, Japan, and Australia (Osborne et al., 2001). Symptoms include itching, rash, burning, blisters, and deep skin erosions that can be very painful (Fujiki et al., 1985). Dermatitis has been mostly attributed to the toxins aplysiatoxin and debromoaplysiatoxin (Osborne et al., 2001). Accidental human oral exposure to L. majuscula after its incidental ingestion with seaweed resulted in a burning sensation and “scalding” in the mouth (Sims & Zandee van Rilland, 1981) that lasted for up to 3 weeks (Marshall & Vogt, 1998). A fatal human intoxication caused by the consumption of meat from a green turtle, Chelonia mydas, was recently associated with lyngbyatoxin-a in Madagascar (Yasumoto, 1998). Nodularia blooms are found in brackish water systems and have been commonly reported from Australia, New Zealand, and the Baltic Sea (Baker & Humpage, 1994; Carmichael et al., 1988; Sivonen et al., 1989). Nodularia spumigena has been associated with livestock, canine, and wildlife mortality events (see summary in Landsberg, 2002), caused mostly by the consumption of nodularin by terrestrial animals drinking from contaminated water sources. Thus far, only N . spumigena has been determined to produce nodularin (Bolch et al., 1999), a cyclic pentapeptide hepatotoxin that is a protein phosphatase 1 and 2A inhibitor, and tumor promoter (Honkanen et al., 1991; Yoshizawa et al., 1990). There are potential human health risks associated with shellfish that have consumed Nodularia. Tissues from edible blue mussels, Mytilus edulis, tested during an N . spumigena bloom in Peel-Harvey Inlet, western Australia, were toxic to mice and induced characteristic hepatic hemorrhage and hepatocyte degeneration. However, only the mussel gastrointestinal tracts retained toxicity after the Nodularia bloom ended. It was recommended that mussels should not be collected for human consumption during a Nodularia bloom (Falconer et al., 1992). When N . spumigena counts reached 4 × 104 cells/ml, there was a sufficient concentration of toxin in edible mussels to post warnings against recreational harvesting. Toxin concentrations in prawns also exceeded health alert levels (Van Buynder et al., 2001). Few studies have associated nodularin with aquatic animal mortalities. Experimental exposures of animals to N . spumigena cultures demonstrated that the toxins induced severe but reversible liver damage in sea trout, Salmo trutta m. trutta (Kankaanp¨aa¨ et al., 2002), affected larval development of herring, Clupea harengus membras, and negatively influenced the survival of copepods, Eurytemora affinis (Ojaveer et al., 2003). Fish and crabs have been reported to avoid the waters of the estuaries during Nodularia blooms (Potter et al., 1983). Nodularin has been found naturally accumulated in marine mussels (M. edulis, Dreissena polymorpha) and clams (Macoma balthica), and in the marine fish, flounder (Platichthys flesus) and cod (Gadus morhua) (Sipi¨a et al., 2001, 2002).

8.4. OTHER HAB SPECIES CAUSING IMPACTS ON LIVING RESOURCES HABs have caused multiple impacts on a range of ecosystems, populations, and species. HABs and their toxins can affect aquatic organisms through a number of mechanisms (see

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Landsberg, 2002). The following discussion pertains primarily to HABs that have no known effects on public health or higher vertebrates, but through the production of a suite of bioactive or toxic compounds can be responsible for the deaths of millions of fish and shellfish globally. Economically, such blooms can cause significant losses to local industries. At least 60 HAB species have been documented to be ichthyotoxic alone (Landsberg, 2002); only a few selected species are discussed here.

8.4.1. Raphidophytes (Chattonella, Fibrocapsa, Heterosigma), Reactive Oxygen Species, Brevetoxins, Fish Kills The raphidophytes Chattonella, Fibrocapsa, and Heterosigma are primarily known to be ichthyotoxic. Mechanisms of toxicity include the production of brevetoxins (see page 171), brevetoxin-like compounds, reactive oxygen species (ROS) (e.g., superoxide anions, hydroxyl radicals, singlet oxygen, and hydrogen peroxide), and hemolytic fatty acids (Bourdelais et al., 2002; Hallegraeff et al., 1998; Khan et al., 1996; Marshall et al., 2003; Oda et al., 1997; Tanaka et al., 1992; Twiner & Trick, 2000). Some fish mortalities may therefore be attributable to the production of any of these bioactive products, either singly or in various combinations. ROS and free fatty acids contribute to pathological changes in the gills of fish (Marshall et al., 2003), while brevetoxin-like neurotoxins impair cardiac function (Endo et al., 1992). Raphidophytes cause significant impacts in natural systems and in aquaculture facilities, resulting in substantial economic losses. Chattonella spp. have caused fish mortalities in the Far East, particularly in Japan. In the summer of 1972, in the Seto Inland Sea, a C. antiqua red tide caused a mass mortality of about fourteen million cultured yellowtail Seriola quinqueradiata, worth more than seven billion yen (Imai et al., 1998). When fish are exposed to Chattonella, pathological damage occurs in the gills, leading to impaired osmoregulation, edema, and reduced oxygen transfer, and resulting in asphyxiation (Endo et al., 1985; Ishimatsu et al., 1996; Shimada et al., 1983). In April–May 1996 a mortality of an estimated 1700 tonnes of cultured southern bluefin tuna, Thunnus maccoyi, in Boston Bay, Australia coincided with a bloom of Chattonella marina. Fish were observed to be distressed, swimming in a haphazard manner on the surface, and in some cases, gasping. Gill pathology included excess mucus production, marked epithelial swelling, and separation of the epithelium. Brevetoxin-like compounds were detected in the tuna livers at up to 142 µg per 100 g of tissue using a sodium-channel receptorbinding assay. A likely scenario for mortality caused by Chattonella spp. is that fish die because of the direct toxic effect of oxygen radicals and free fatty acids on the gills in combination with the cardiotoxic effects of brevetoxins (Hallegraeff et al., 1998; Ishimatsu et al., 1990; Marshall et al., 2003). In July–September 2000, brevetoxins produced by Chattonella cf. verruculosa were implicated in the USA for the first time in mortalities of 1–2.5 million Atlantic menhaden, Brevoortia tyrannus, in Delaware (Bourdelais et al., 2002). In the Seto Inland Sea, Japan, the economic damage caused by Heterosigma akashiwo red tides totaled 2 billion yen from 1972 to 1987 (Honjo, 1993). Blooms of H. akashiwo have also been associated with fish kills in British Columbia since 1976. In 1986, Heterosigma blooms led to losses to the salmonid industry of at least $2.5 million, which at the time represented a third of the value normally produced (Taylor, 1993). Penned salmonids die when concentrations of H. akashiwo reach several million cells per liter, and some species—such as rainbow trout, Oncorhynchus mykiss, and Atlantic salmon, Salmo salar—are more sensitive to Heterosigma than others, such as chinook salmon, Oncorhynchus tshawytscha (Taylor & Haigh, 1993). In

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early January 1989, a dense bloom of H. akashiwo was associated with chinook salmon kills in New Zealand. Affected salmon showed degenerative changes of the branchial epithelium and vasculature, including swelling of the respiratory epithelium, mucus discharge, and separation of the secondary epithelium from the pillar cells by edema. Impairment of respiratory and osmoregulatory functions of the gills was concluded to be the cause of death (Chang et al., 1990). H. akashiwo generates reactive oxygen intermediates that probably first attack epithelial cell membranes of the secondary gill lamellae, causing lipid peroxidation damage, inactivation of ATPase of the epithelial cell plasma membrane, and edema. When H. akashiwo cultures were disturbed, the release of superoxide increased, thus mimicking the situation that occurs when water containing H. akashiwo passes over the secondary lamellae of finfish (Yang et al., 1995).

8.4.2. Prymenesiophytes (Prymnesium, Chrysochromulina), Prymnesins, Fish Kills At least four species of Prymnesium are currently recognized as being toxic to aquatic organisms. Ichthyotoxic Prymnesium blooms have been documented since the late 1800s, mostly from brackish-water systems in Israel, Europe, the Ukraine, China, North America, and Australia. Recreational fisheries, particularly of largemouth bass, in Texas, USA have been severely affected in the last few years with economic losses estimated to be more than $7 million (Texas Parks and Wildlife Department, 2003). In particular, Prymnesium parvum has been responsible for significant global economic losses caused by fish kills in aquaculture facilities and fish farms (Moestrup, 1994; Sarig, 1971). Except for the marine species P. calathiferum (Chang & Ryan, 1985), Prymnesium blooms occur in low-salinity (usually between 1 and 12 ppt), brackish waters in ponds, shallow lakes, or lagoons (Edvardsen & Paasche, 1998). Recurrent kills in brackish fish ponds in Israel have been a significant problem since the early 1940s, where under suitable conditions, blooms of P. parvum can become dominant within 3–5 days (Sarig, 1971). Routine fish bioassays using suspect pond water and mosquito fish, Gambusia affinis, have helped fish farmers take preventative measures and avoid unnecessary economic losses. Ponds are chemically treated with copper or ammonium sulfate when low concentrations of P. parvum are confirmed toxic by the fish bioassay (Sarig, 1971). Fish kills usually occur when Prymnesium cell densities exceed 5–10 × 105 cells/ml. Since 1989, in the Ryfylke fjords, Norway, annual mixed blooms of P. parvum and Prymnesium patelliferum have caused extensive damage to farmed fish (Edvardsen & Paasche, 1998). The biological spectrum of activity of prymnesins is extremely broad (Igarashi et al., 1995; Shilo, 1981). Cultures of both P. parvum and P. patelliferum were toxic to Artemia, and crude lipid extracts had toxic effects on human erythrocytes and synaptic vesicles of rat brains (Meldahl et al., 1994; Meldahl & Fonnum, 1995). At least, two toxins, prymnesin-1 and prymnesin-2, with hemolytic and ichthyotoxic properties have been purified (Igarashi et al., 1995, 1996). Prymnesins act on the cytoplasmic membranes of gill-breathing animals, such as fish, tadpoles, and molluscs, which are particularly sensitive to toxins because of the increased permeability of their gill membranes, especially in conditions that are cation-activated, pHdependent, and sodium chloride-inhibited (Shilo, 1981; Terao et al., 1996). Ichthyotoxicity is enhanced in the presence of cations, which increase the permeability of the gills to toxin uptake (Ulitzur & Shilo, 1964). Killifish immersed in a 2-ppm prymnesin solution died at a

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lower concentration (0.5–1.0 ppm) when calcium ions (0.2–0.5%) were added to the solution (Igarashi et al., 1995). Natural fish kills associated with P. parvum in Texas, USA are correlated with the distribution of rivers and reservoirs overlying calcium deposits (Glass, Texas Parks and Wildlife Department, personal communication), demonstrating the likely connection between increased Prymnesium toxicity and cationic water chemistry.

8.4.3. Dinoflagellates (Cochlodinium), Reactive Oxygen Species, Fish Kills The dinoflagellates Cochlodinium spp. are another group of microalgae producing bioactive ROS, hemolysins, and polysaccharides (Kim et al., 1999, 2002; Lee, 1996; Yuki & Yoshimatsu, 1989), which are only known to affect invertebrates and lower vertebrates such as fish (Guzm´an et al., 1990; Qi et al., 1993; Yuki & Yoshimatsu, 1989). Cochlodinium spp. are increasingly causing problems to the mariculture and fisheries industries, resulting thus far in maximum annual losses of $95.5 million in Korea (Kim, 1998; Kim et al., 1999) and CDN $2 million in the Pacific Northwest (Whyte et al., 2001). In the Far East, culturists are actively investigating control methods using loess on a large scale (Kim, 1998). In late August 1985, a small-scale Cochlodinium sp. red tide occurred in Harima-Nada, Japan, where cultured yellowtails, Seriola quinqueradiata, showed unusual behavior, possibly because of toxin exposure (Yuki & Yoshimatsu, 1989). During mid-June 1990, an unusual red tide event, a toxic bloom of Cochlodinium sp., occurred from Weitou Bay to Quanzhou Bay, China. A wide range of fish, shellfish, and jellyfish species were killed. Crustaceans such as crabs, lobsters, and shrimp were not affected (Qi et al., 1993). In the Pacific Northwest, Cochlodinium sp. blooms caused substantial mortalities of farmed Atlantic salmon, S. salar, and chinook salmon, Oncorhynchus kisutch. When cells exceeded 5 × 102 cells/ml fish stopped feeding and at concentrations of more than 2 × 103 cells/ml mortalities occurred. Experimental field bioassays with salmon smolts demonstrated lethality after 120 m exposure with over 90% mortality after 500 m when cell concentrations varied from 2.7 to 10.8 × 103 cells/ml. In fish bioassays at densities of above 1 × 104 cells/ml, fish showed signs of respiratory distress, loss in equilibrium, and an inability to stay in the water column (Whyte et al., 2001).

8.4.4. Dinoflagellates (Heterocapsa), Uncharacterized, Shellfish Mortalities In September 1988, a harmful bloom of the dinoflagellate Heterocapsa circularisquama (Horiguchi, 1995) first recorded in Uranouchi Bay, Kochi Prefecture, Japan was associated with bivalve mortalities. The following year this species was responsible for the mass mortality of Japanese littleneck clams (Venerupis philippinarum), Mediterranean mussels (Mytilus galloprovincialis), Pacific oysters (Crassostrea gigas), razor clams (Solen strictus), and surfclams (Mactra chinensis) in Fukuoka Bay. Subsequently H. circularisquama has been implicated in other mass mortalities of bivalves. In the last decade, economic losses to the shellfish aquaculture industry from the direct killing of marketable products by H. circularisquama were estimated at 10 billion yen (= approx. $1 million) (Matsuyama, 1999). Laboratory exposure studies have demonstrated that in addition to bivalves, organisms such as gastropods, solitary ascidians, jellyfish, and protists (including dinoflagellates) are also affected by H. circularisquama. No demonstrable effect on mice, finfish, crabs, lobsters, shrimp, starfish, copepods, or diatoms could be found (Matsuyama, 1999). Detrimental effects to wild or cultured populations of fish or crustaceans or to public health have not been

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observed (Matsuyama et al., 1996, 1997). The toxin of H. circularisquama has still not been characterized. Toxicity appears to be closely related to the outer components of the cells, and a toxic, proteinaceous substance is probably present on the cell surface (Matsuyama et al., 1996, 1997). Purification and characterization of toxic fractions have not been successful because of their high lability under neutral conditions (Matsuyama, 1999). Recent studies demonstrate that bioactivity of H. circularisquama cells results in hemolysis of mammalian erythrocytes. Oda et al. (2001) suggested that a hemolytic toxin may contribute to the shellfish mortalities associated with H. circularisquama.

8.5. MODES OF EXPOSURE Typically, aquatic animals are exposed to harmful algae when planktonic or epibenthic species bloom and dominate the food web, but there are also less obvious means of exposure. Whether HAB species are benthic or planktonic, predatory or photosynthetic, or in a resting cyst phase will influence how organisms are exposed. The degree of harm incurred will depend on whether susceptible organisms can detect blooms or toxins and then avoid them. In some cases, there may be no obvious mechanism by which organisms can detect toxins, and so they may inadvertently consume toxic HAB species or prey. Organisms are directly exposed to microalgal cells and their toxins by direct contact, or either by drinking them or ingesting them via various feeding modes (e.g. filter feeding, predation). Filter-feeding zooplankton, sponges, and shellfish can take up toxic cells directly from the water column and then accumulate the toxins in the viscera. Planktivorous fish that actively prey upon toxic microalgae can also retain toxins. Because filter feeders and predators are not necessarily discriminatory, they can be exposed to most of the known major microalgal toxins. Numerous toxic microalgae have sedimentary cyst or resting stages as part of their life cycle. In some cases, these stages may be more toxic than their free-swimming planktonic counterparts. When newly formed, Alexandrium cysts can be up to 1000 times more toxic than the vegetative cells (Dale & Yentsch, 1978). Otherwise, toxic cells or adsorbed toxins may sink to the bottom, be consumed by benthic organisms, and then recycled into the food chain. During active growth many microalgae release exudates such as extracellular toxins (exotoxins) into the surrounding water; organisms can be affected by these products even in the apparent absence of cells. Release of exotoxins by microalgae and the long-term persistence of toxins in water are determined by basic physicochemical properties that influence their stability. Release of toxins into a watery milieu may not always pose a threat to organisms if large volumes of water dilute the toxins. Some organisms, such as zooplankton, are able to detect toxins via chemoreception. For example, soluble saxitoxins acting as feeding deterrents can be detected by copepods (Shaw et al., 1997). Organisms can be affected by direct cell-to-cell contact (Uchida et al., 1995) with microalgae, either through exposure to microalgal toxins present on the cell surface or through the mechanical damage caused when microalgal anatomical structures penetrate the gills or skin of the exposed organism (see page 186). Many species that produce intracellular toxins generally release little into the surrounding water under normal conditions. Under stressful environmental conditions (e.g., salinity change, wind action or currents, or during the termination phases of a bloom), algae release toxins frequently as a result of lysing cells. However, a decline in physiological status (in the absence of cell lysis) may also promote toxin release, as in the case

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of high dissolved domoic acid levels (>120 nM) associated with a dense Pseudo-nitzschia australis population in Monterey Bay, CA (>2 × 106 cells/l), which appeared to have been at or approaching senescence (Doucette et al., 2002b; Scholin, unpublished data). In terms of cell disruption, the relative structure of the microalgal cell—armored, lightly armored, or unarmored—may determine its resistance to lysis and will influence the release of toxins into the water. Treating blooms with chemicals or by manipulating the environment can result in high concentrations of toxins in the water that are more damaging to aquatic organisms than the bloom itself. Several planktonic species can release toxins that become aerosolized upon lysis or that become caught up in bubble-mediated transport. Bubble-mediated transport has been shown to concentrate brevetoxins from K. brevis at the sea surface, where concentrated toxins are subsequently released as an aerosol (Pierce et al., 1990). Air-breathing mammals, including humans, and reptiles can be adversely affected by aerosolized toxins. Brevetoxins can be absorbed directly across the gill membranes of fish (Abbott et al., 1975) although ingestion through the food chain is another documented route (Tester et al., 2000). Fish kills usually originate with the lysis of toxic K. brevis cells and the subsequent liberation of toxin and passage across the gills (Baden, 1988). Biotoxins are transferred trophically when organisms consume other organisms that have been directly exposed to toxic microalgae and have bioaccumulated, bioconverted, or biomagnified the toxins. In many cases, the organism that was consumed has modified the toxins from the form that was originally produced by the microalgae. Filter-feeding zooplankton consume and retain toxic microalgae that are then passed one step up the food chain to zooplanktivores (e.g., Turner et al., 2000), which can range from planktivorous fish to cetaceans. Filter-feeding shellfish also consume toxic microalgae and accumulate the toxins, which in turn become available to both animal and human consumers. This transfer of toxins up the food chain is one of the most common ways in which higher trophic levels, including humans, are affected by microalgal toxins. People who consume toxic tropical fish and become sick with CFP are victims of indirect exposure to microalgal toxins. Some ciguatoxins are biomagnified and some are biotransformed during their transfer up the food chain from toxic epiphytic dinoflagellates ingested by herbivores, to piscivores that feed on the toxic herbivores, and finally to humans consuming toxic piscivorous fish (Swift & Swift, 1993). The toxins found in benthic dinoflagellates are transformed into a form available to and harmful to humans by the time they are concentrated in piscivorous fish (see page 174).

8.6. EMERGING ISSUES 8.6.1. Indirect Effects Harmful algal blooms can have a range of indirect effects upon ecosystems, habitats, and upon species. HABs can cause mechanical damage and affect physiological processes or lead to direct pathologies. Organisms can be affected when anatomical structures penetrate the gills or skin. Some diatoms, for example, Chaetoceros spp., Skeletonema costatum, and Rhizosolenia chunii have setae, barbs, processes, or spines that can become trapped in the gills of fish or shellfish, inflict mechanical damage that impairs respiration, and eventually cause death (Kent et al., 1995; Parry et al., 1989; Tester & Mahoney, 1995). Mortalities of Atlantic salmon, S. salar, reared in net pens in British Columbia, Canada, were indirectly associated with a dense bloom of mixed diatoms dominated mostly by S. costatum and Thalassiosira species (Kent

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et al., 1995). Chaetoceros species have also been considered responsible for mortalities of salmon in Scotland (Bruno et al., 1989) and in the Pacific Northwest (Rensel, 1993; Yang & Albright, 1994), and of red king crabs, Paralithodes camtschatica, in Alaska (Tester & Mahoney, 1995). Mortalities of cultured fish in Europe have been attributed to Dictyocha speculum blooms (Erard-Le Denn & Ryckaert, 1990; Henriksen et al., 1993). Penetration or irritation of the fish by the microalgal spines can cause a chain of events in the gills; excess mucus production, clogging, reduced oxygen uptake, hemorrhaging, blood hypoxia, occasional secondary bacterial infection, or lesions. These effects, coupled with likely low levels of dissolved oxygen due to the blooms, usually resulted in fish death due to asphyxiation. Any highly concentrated bloom has the potential to reduce water quality, being affected by HAB metabolic activities or by their mechanical presence. Decomposing and respiring cells can contribute to the reduction (hypoxia) or absence of oxygen (anoxia), production of ammonia, and other nitrogenous by-products, or to the formation of toxic sulfides. Declining water quality can lead to mortalities or chronic disease conditions, avoidance of species in the area, and reduced feeding. Such sublethal, chronic effects on habitats can have far-reaching impacts on animal and plant communities. Significant economic losses from mass mortalities of aquatic organisms have been blamed on poor water quality conditions caused by blooms of Ceratium (Mahoney & Steimle, 1979; Onoue, 1990; Pitcher & Cockcroft, 1998), mixed Ceratium and Prorocentrum micans (Matthews & Pitcher, 1996), Gonyaulax polygramma (Grindley & Taylor, 1964; Koizumi et al., 1996; Lam & Yip, 1990), Noctiluca (Adnan, 1989; Elbr¨achter & Qi, 1998; La Barbera Sanchez, 1991; Subramanian, 1985), and Akashiwo sanguinea (Harper & Guillen, 1989). The large biomass and long-term persistence of brown tides caused by Aureococcus anophagefferens and Aureoumbra lagunensis result in numerous direct and indirect effects on aquatic organisms and their habitats. Dramatic losses in species abundance and diversity have been attributed to inhibition of zooplankton grazing or filter feeding in shellfish, starvation and recruitment failure, and chronic toxicity (Buskey et al., 1996; Bricelj & Lonsdale, 1997; Cosper et al., 1987; Dennison et al., 1989; Tracey, 1988). After direct contact with the brown tide pelagophyte Aureococcus anophagefferens, adult blue mussels, Mytilus edulis (Tracey, 1988), and larval bay scallops, Argopecten irradians (Gallager et al., 1989), were inhibited from feeding (Bricelj & Kuenstner, 1989). In 1985 and 1986, A. anophagefferens was responsible for the demise of the bay scallop fishery on Long Island, New York, USA, by inducing starvation and subsequent recruitment failure (Tettelbach & Wenczel, 1993). Economic losses to the bay scallop fishery of New York State caused by the reduced landings attributed to brown tide were estimated at $2 million per year (Kahn & Rockel, 1988). Persistent brown tides reduced light availability, which resulted in a decline in eelgrass, Zostera marina (Dennison et al., 1989). Because eelgrass provides shelter and substrate for numerous estuarine benthic species, such losses in habitat may have contributed to the poor recruitment of bay scallops (Tettelbach & Wenczel, 1993). In Laguna Madre, Texas, USA, a persistent, nearly monospecific bloom of Aureoumbra lagunensis has been responsible for declines in meso- and micro-zooplankton populations; surfclams, Mulinia lateralis; larval fish, and seagrass (Buskey & Hyatt, 1995).

8.6.2. Sublethal, Chronic, and Repeated Exposures Studies on the effects of algal toxins on humans and animals are largely focused on seafood poisoning incidents or mass mortality events associated with their acute neurotoxic activity. In

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contrast, the effects of sublethal, chronic, and repeat exposures remain poorly documented. Experimental evidence suggests the involvement of certain algal toxins in developmental toxicity, immunosuppression, and tumorigenicity; however, none of these effects have been demonstrated definitively in the environment. Exposure to sublethal concentrations of many toxins can lead to behavioral changes that result in reduced fitness. For example, sublethal responses of mollusks, crustaceans, and zooplankton to toxins and toxic algal species generally include impaired feeding, which leads to reduced growth, and consequently decreased reproduction (reviewed by Landsberg, 2002). At a higher trophic level, saxitoxins are suspected to play a role in decreased reproductive success of endangered North Atlantic right whales (Eubalaena glacialis), which encounter saxitoxin-producing Alexandrium fundyense blooms almost annually in their summer feeding grounds in the Bay of Fundy. The copepod Calanus finmarchicus dominates the right whale’s diet in the lower Bay of Fundy, and is known to feed upon A. fundyense. Based on an ingestion rate of 4 × 108 copepods/day, it is estimated that right whales may ingest between 4.73 and 9.65 µg STX/kg/day during blooms (Durbin et al., 2002). This is similar to the estimated lethal dose in humans (Levin, 1992). Saxitoxin measured in right whale feces from the area can exceed 20 µg/g (Doucette et al., 2002a). The significance of this toxin concentration is not yet clear with respect to the level of saxitoxins circulating in the whale’s blood, or with regard to its effects on behavior and physiology. However, sublethal exposure to saxitoxin is proposed to affect swimming and diving behavior, which in turn could result in decreased feeding rates, poorer fitness, and ultimately reduced calving rates (Durbin et al. 2002). Consistent with this hypothesis, the diving frequency of right whales in the Bay of Fundy is reported to be lower than that of right whales in Cape Cod Bay and their breathing rate is higher (Durbin et al. 2002). Thus, the chronic effects of saxitoxins on right whale reproductive fitness warrants further investigation. A potential for an increased susceptibility to other impacts on the right whale population (e.g., ship strikes) as a result of compromised swimming/diving abilities due to toxin exposure is also an area requiring additional study. In a study of the effects of repeated low doses of domoic acid in rats, Peng et al. (1997) found no additive effects of domoic acid on learning and memory processing at doses below those known to cause hippocampal lesions. Because domoic acid is efficiently cleared from the circulation by the kidney, it appears that it is only when the clearance mechanisms are overwhelmed that neurotoxic consequences occur. In humans, long-term consequences of sublethal exposure to domoic acid are manifested in persistent short-term memory loss (Perl et al., 1990), most likely as a consequence of hippocampal lesions in the brain. Therefore, long-term consequences of sublethal domoic acid poisoning in marine animals might also be expected. Indeed, episodic seizures have been observed in tagged sea lions released after rehabilitation following domoic acid poisoning associated with Pseudo-nitzschia blooms on the California coast, USA (Gulland, unpublished data). The long-term survival of released animals has not been documented. Domoic acid has been identified as a cause of reproductive failure in sea lions, which may have population level consequences if blooms of Pseudonitzschia on the California coast continue on a near annual basis as they have since 1998. In 2003, a domoic acid event in the area of the Channel Islands, the major pupping grounds for California sea lions, resulted in nearly 100% mortality of the 2003 year pups, due to still births, premature parturition associated with maternal seizures, and/or abnormal pup development (Gulland et al., unpublished data). Domoic acid appears to be transferred across the placenta, as it is measurable in fetal fluids. Whether abnormal pup development is a direct

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consequence of domoic acid exposure or the indirect consequence of maternal seizures is unknown. Developmental toxicity has been demonstrated in fish embryos in response to ciguatoxins, which are known to bioaccumulate in the viscera and flesh of higher trophic level fish (page 174). Because lipophilic contaminants are often transferred to offspring during oogenesis, Edmunds et al. (1999) used an egg microinjection paradigm that parallels maternal transfer of toxins, to demonstrate cardiovascular, muscular, and skeletal abnormalities and reduced hatching success of medaka eggs. They suggested that the sensitivity of embryonic fish to ciguatoxin via direct oocyte exposure may indicate that maternal transfer of low levels of ciguatoxin is an unrecognized threat to the reproductive success of tropical reef fish. Some similar, as well as distinctly different, developmental abnormalities were observed in medaka embryos in response to the brevetoxins, PbTx-1 (Kimm-Brinson & Ramsdell, 2001) or PbTx-3 (Coleman & Ramsdell, 2002). The mechanisms by which these three toxins, all of which target site 5 on voltagedependent sodium channels, cause different developmental defects remain to be explored. Although acute exposure to lethal doses of brevetoxin results in massive animal mortalities, effects of low-level exposure to brevetoxins are harder to interpret. For example, in southwest Florida, USA, manatees are routinely exposed to brevetoxins from K. brevis red tides (see page 173). Bossart et al. (1998) postulated that chronic exposure to aerosolized brevetoxins could affect the cellular immune system of manatees and compromise their response to disease or toxin exposure. In addition to their actions on voltage-dependent sodium channels, indirect evidence suggests that brevetoxins may cause immune suppression by inhibiting the activity of cysteine cathepsins, a class of lysosomal proteases that function in antigen presentation in B-cells (Baden, 1996; Bossart et al., 1998; Sudarsanam et al., 1992). Consistent with this hypothesis, Bossart et al. (1998), using an immunohistochemical staining procedure, confirmed that brevetoxin was localized in manatee lungs. Colocalization of interleukin-1b-converting enzyme (ICE) in the lung tissues identified the brevetoxin-containing cells as activated macrophages. ICE is a key mediator in the release of inflammatory cytokines from activated macrophages, the release of which can result in fatal toxic shock (Bossart et al., 1998). Thus, while brevetoxin does appear to interact specifically with immune cells in these tissues, immunosuppression by brevetoxin has not yet been confirmed. The okadaic acid family of polyether toxins is unique among the algal toxins, in that it targets serine/threonine protein phosphatases involved ubiquitously in intracellular signaling mechanisms that are frequently altered in tumor formation. Okadaic acid has been shown to induce skin papillomas and carcinomas in mice and adenocarcinomas in rats (Fujiki et al., 1988; Suganuma et al., 1990; Fujiki & Suganuma, 1993). However, studies on the chronic effects of exposure to okadaic acid or dinophysistoxins have not been conducted in humans. An epidemiological study of digestive-tract cancer mortality in relation to the distribution of DSP was recently conducted in France and although there appeared to be a very tentative positive association between the two, Cordier et al. (2000) recognized the need for more extensive surveys and in-depth research before any link can be definitively proven. Correlation between exposure to dinoflagellates producing okadaic acid and the occurrence of fibropapillomatosis in marine turtles has been demonstrated, but a causal linkage has not been made (Landsberg et al., 1999). Lyngbya majuscula produces antineoplastic compounds (Luesch et al., 2002) and the tumor-promoters, aplysiatoxin, debromoaplysiatoxin, bromoaplysiatoxin, and lyngbyatoxins (Fujiki et al., 1985; Fujiki & Suganuma, 1996). Lyngbyatoxin-a has been experimentally

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demonstrated to induce papillomas in two-stage mouse carcinogenesis experiments (Fujiki et al., 1984), but unlike okadaic acid, and nodularin, lyngbyatoxins promote tumor growth through the activation of protein kinase C, not through protein phosphatase inhibition (Fujiki & Suganuma, 1996). That lyngbyatoxin-a was recently identified in green turtle meat (Yasumoto, 1998) confirms that these toxins, like okadaic acid, should also be considered for their potential role as tumor-promoters in sea turtle fibropapillomatosis (Landsberg et al., 1999).

8.6.3. HABs as Potential Vectors for Pathogens and as Stressors for Disease Apart from the intrinsic toxicity of many harmful algal species, scientists have begun to examine more closely the possibility that blooms of these organisms may harbor bacterial or viral pathogens. Many of the conditions that are favorable for the initiation and development of HABs are also conducive to the maintenance and growth of pathogens (Henrickson et al., 2001). There is also increasing evidence that both bacteria and viruses causing disease in humans can remain viable longer in the marine, and especially the coastal/estuarine, environment than previously thought (Birch & Gust, 1989; Colwell & Spira, 1992). Moreover, an indirect link between coastal algal blooms and cholera epidemics has been proposed (Epstein, 1995). While such a scenario is consistent with reports of close associations between planktonic species and human pathogenic organisms (Huq et al., 1995), the suggestion that algal blooms serve as effective reservoirs and vectors for such pathogens remains speculative. By comparison, there is a strong likelihood that HABs may harbor and facilitate the transmission of animal (e.g., fish, shellfish) pathogens in aquatic systems. The increased susceptibility of aquatic organisms to bacterial disease during or after algal blooms has been well documented and a number of well-known pathogens, including Vibrio alginolyticus, V. anguillarum, V. parahaemolyticus, and Aeromonas hydrophila, have been isolated in association with ongoing HAB events (see review by Landsberg, 2002). Indeed, by compromising the health of animals through physically (e.g., spines, polysaccharides) or chemically (e.g., toxins, hypoxia) mediated stress, as well as providing an organically rich growth environment for bacteria, HAB species could effectively promote pathogenic infections. Some aquatic mortality and disease events may therefore be strongly interrelated with pathogenic microbe and HAB associations, instead of separately triggered by either toxic microalgae or infectious pathogens, the currently accepted model (Landsberg, 2002). Obtaining definitive evidence that those bacteria responsible for diseased animals are the same species vectored by co-occurring algal populations will be the aim of future investigations. Although not well understood or typically linked as having a role in marine animal morbidity and disease syndromes, HABs and their toxins are emerging as significant factors (Landsberg, 2002). Recent examples drawn from freshwater–terrestrial systems highlight the need for animal disease investigations to seriously consider apparently unrelated HAB events as possible etiological factors in animal mortalities. For example, there has been some discussion about the potential role of the cyanobacterial microcystins contributing as stressors in avian botulism (Murphy et al., 2000) and possible links between the freshwater cyanobacterium Hapalosiphon and avian vacuolar myelinopathy, a neurological disease affecting aquatic birds and bald eagles in the southeastern USA (Wilde et al., 2003). Exposure to microalgal cells or their toxins may also be sufficient to cause disease, or to increase the susceptibility of aquatic organisms to opportunistic pathogens. For example, during the 1990s, more than 20 species of tropical fish along Florida’s coral reef tract in the USA were involved in a chronic mass mortality. Fish were affected by multiple opportunistic

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protistan parasites and bacterial pathogens considered to be secondary invaders of fish whose health had already been compromised. There was strong circumstantial evidence that fish could have been chronically affected and immunocompromised by exposure to toxins from benthic dinoflagellates, such as those involved in CFP (Landsberg, 1995) (see page 174). In Australia, planktonic blooms dominated by the cyanobacterium Oscillatoria corakiana were coincident with mortalities of cultured prawns, Penaeus monodon. Vibrio species (including V. alginolyticus, V. harveyi, V. vulnificus, V. parahaemolyticus, V. anguillarum) were in higher abundance in tissues of morbid prawns compared to those of healthy prawns. Sublethal levels of Oscillatoria toxin(s) weakened the prawns, possibly via a neurotoxic effect that reduced feeding behavior and impaired the immune system, and prawns were susceptible to secondary bacterial infections (Smith, 1996). During 1989–1990, a Noctiluca scintillans bloom in the Northern China Sea resulted in losses of about $100 million to shrimp (Penaeus orientalis) mariculture operations. Poor water quality caused an increase in the susceptibility of prawns to parasitic infections, which led to disease, mortalities, and additional losses valued at $1 million (Chen & Gu, 1993). In August–September 2001, Kuwait Bay, a semi-enclosed embayment of the Arabian Gulf, experienced a massive fish kill involving more than 2500 metric tonnes of wild mullet (Liza klunzingeri) due to the bacterium Streptococcus agalactiae. The event was preceded by a small fish kill of gilthead seabream in aquaculture net pens that was associated with a bloom of the dinoflagellate Ceratium furca. Unusually warm temperatures and calm conditions, a succession of harmful algal bloom species (e.g., Karenia selliformis Fig. 8.8), fluctuating nutrient inputs

Figure 8.8. Toxic Marine dinoflagellate. Karenia selliformis, bar = 10 µm.

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from localized sewage outfalls, and decomposing fish triggered by the initial HAB event were considered major factors that contributed to fish stress and subsequent disease (Glibert et al., 2002). Microalgae with external structures or that produce bioactive compounds have a potential role in aquatic animal diseases if they are capable of damaging external surfaces and providing a portal of entry for pathogens. In addition to mechanically damaging gill surfaces (see page 186), diatoms have been associated with increasing the susceptibility of fish to disease. Cultured salmonids were more susceptible to vibriosis and/or bacterial kidney disease when exposed to low concentrations of Chaetoceros concavicornis and Chaetoceros convolutus (Albright et al., 1993; Speare et al., 1989). Fish exposed in aquaria directly to Pfiesteria piscicida cells or to water filtrate (and presumptive toxin) that contained P. piscicida showed behavioral changes and sloughing of the fish epithelium, which led to invasion by numerous secondary, opportunistic pathogens (Noga et al., 1996). That Pfiesteria causes mechanical damage to fish skin during pedunculate feeding, which could allow for subsequent invasion by secondary opportunistic pathogens, is a possible mode of lesion development (Vogelbein et al., 2001).

8.7. ANTHROPOGENIC INPUTS AND STRESSORS There is increasing evidence and concern that the introduction and global expansion of HAB species is partially caused by anthropogenic influences (Hallegraeff, 1993). However, increases related to man-made activities should be distinguished from natural stressors and perturbations that can also contribute to HAB events or can act synergistically with anthropogenic risks. Anthropogenic factors can cause an increase in the frequency, persistence, species domination, or expansion in the geographic range of HABs and provide appropriate environmental conditions conducive to increased toxin production. Harmful algae can respond to a variety of anthropogenic inputs that include nutrients and minerals by increasing their biomass. Land-based (industrial, agricultural, and water management) and aquaculture activities can negatively influence coastal conditions causing eutrophication and fluctuating salinities. Phytoplankton productivity in coastal waters is controlled primarily by the availability of nitrogen (Paerl, 1997). Anthropogenic nitrogen sources, primarily agriculture and aquaculture, fossil fuel burning, and sewage and animal waste, have led to an increase in riverine input of nitrogen into the North Atlantic by 2–20 fold since preindustrial times (Howarth, 1998) and overall, coastal waters in developed countries have experienced a long-term increase in the loading of both nitrogen and phosphorus by more than a factor of 4 (Nixon, 1995). A range of landbased or atmospherically derived inputs such as nitrogen and nitrogenous products (ammonia, urea), phosphorus, iron, vitamins, minerals, and humics (Anderson et al., 2002; Bates et al., 2001; Maldonado et al., 2002; Paerl, 1997) has been positively demonstrated to increase algal biomass or toxicity and contribute to the formation of problematic HABs in coastal waters. Not only is the total concentration of enhanced nutrients of significance, but the consequences of altered ratios of nutrients are also relevant. Increases in nitrogen (N) and phosphorus (P) can result in silica (Si) limitation due to decreases in the Si:N and Si:P ratios, favoring the growth of flagellates over silica-dependent diatoms, which tend to be nontoxic (Humbourg et al., 1997; Reigman, 1998).

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Evidence for an increase in HAB biomass or toxicity in relation to anthropogenic inputs is also indicated by anecdotal reports, historical records, sediment cyst profiles, and demonstrated in vitro experimental responses of microalgal species with environmental manipulations (Johannson, & Gran´eli, 1999; Legrand et al., 2001; Parsons et al., 2002; Sohet et al., 1995). Environmental factors can trigger an increase in toxic HABs. For example, that Pseudo-nitzschia respond to increased nutrient enrichment and fluvial flow (Trainer et al., 2001) was demonstrated by the appearance of toxic blooms following heavy rainfall and freshwater nutrient runoff in eastern Canada in 1987 just prior to the ASP incident (Bates et al., 1998). Aquaculture and agriculture activities in coastal systems can have a number of direct or indirect effects on HABs. The presence of high density masses of fish or shellfish can introduce organics and stimulate blooms, stressing fish, causing kills, and providing increased nutrient loading back into the system which in turn can fuel further HAB events (Glibert et al., 2002). There is also compelling evidence that the occurrence of Pfiesteria spp. at toxic levels coincides with the eutrophication of coastal waters through intensive swine and poultry agriculture in areas of North Carolina and Maryland, USA (Paerl, 1997; Boesch et al., 2001; Burkholder et al., 2001). Direct or indirect human-related activities have been partly responsible for the expanded distribution of HAB species in previously unreported areas, leading to increased public health risk and animal mortality events. Mechanisms for HAB introductions include the transfer of resistant algal cyst stages in ballast water, as well as in shipments of live or dead consumer products where stages are either attached to the external surfaces of animals present in the holding water used for transfer, or remain viable in the intestinal tracts. The introduction of HAB species into new areas via ship’s ballast raised awareness of this activity as a major threat to the global transmission of HAB species (Baldwin, 1992; Hallegraeff, 1993, 1998; Macdonald & Davidson, 1998). Numerous records of HAB species from previously unreported areas have been documented in the past few years (Hallegraeff, 1998). PSP (see page 167) was unknown in Australia prior to the 1980s when the first outbreaks showed up in Hobart, Melbourne, and Adelaide associated with three different saxitoxinproducing species, Gymnodinium catenatum, Alexandrium catenella, and A. minutum respectively. Based on anecdotal evidence, historical sampling of cysts in sediments, and confirmation of G. catenatum in ships’ ballast water, it was concluded that this species was introduced in Australia (Hallegraeff, 1993; Hallegraeff & Bolch, 1992). Molecular genetic analyses of different strains and species from geographically separated areas have shown close or identical relationships, thus providing further evidence for introductions. Viable cysts have been demonstrated in the holds of cargo ships that had recently arrived from foreign countries (Hallegraeff & Bolch, 1992; Macdonald & Davidson, 1998). One single ballast tank was estimated to contain more than 300 million toxic dinoflagellate cysts that could be germinated into confirmed toxic cultures (Hallegraeff & Bolch, 1991, 1992). In 1991, the Marine Environmental Protection Committee of the International Maritime Organization adopted voluntary guidelines for ballast water handling by cargo vessels (Hallegraeff, 1993). Protocols utilizing various treatment options and management strategies are being developed to minimize the spread of HABs in ballast water (Hallegraeff, 1998). As well as through waterborne activity and via transportation, animals can also transfer or vector HABs. Invasive species can equally vector toxic HAB species with resistant stages. Although there is a strong awareness for the potential of disease or human pathogen transfer via shellfish, and more stringent controls have been implemented, for example through guidelines

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recommended by the National Shellfish Sanitation Program in the USA, little attention has been paid to the potential risk of transfer of HAB species via animal transportation activities (Shumway, personal communication). As an example, the harmful dinoflagellate Heterocapsa circularisquamata has been causing mass mortalities of shellfish in Japan since 1988 (see page 184). Imada et al. (2001) demonstrated that H. circularisquamata can become established in new areas by transferring in shellfish consignments and that it can expand rapidly in the coastal areas of western Japan. Even if species do not produce resistant cyst stages, they can still produce temporary resting stages that are able to pass through the intestinal tract and remain viable during transportation into new areas previously unaffected by HABs.

8.8. METHODS FOR DETECTION The detection of harmful algal species and the toxins they produce is a rapidly advancing field due, in large part, to the requirement for such data by those charged with the management and mitigation of HABs and their frequently wide-ranging impacts. Moreover, it is becoming clear that the toxicity of harmful algal species can vary widely in natural populations (e.g., Scholin et al., 1999). It is therefore insufficient to focus only on detecting the potentially toxic cells, which may contain sufficient toxin to pose a threat to living resources and public health. Hence, development of technologies for the concurrent detection of organisms and toxins is the ultimate aim of several ongoing efforts (e.g., Scholin et al., 2004) and, where feasible, the incorporation of automated, in situ methods is being pursued. The latter will yield real or near-real-time data that can serve as an early warning to managers of an impending HAB event and aid in the forecasting of bloom development and/or movement when coupled with the appropriate biophysical models. The following discussion briefly outlines some of the current approaches being pursued for the detection of HAB species and toxins, which are the subjects of several recent, detailed reviews (see Botana, 2000; Cembella et al., 2003; Fern´andez et al., 2003; Scholin et al., 2003, 2004; Sellner et al., 2003).

8.8.1. Harmful Algal Species The traditional method used to detect and enumerate harmful species is direct observation of live or preserved material by light microscopy (see Sournia, 1978). This approach provides important morphological confirmation of a species’ presence and abundance. Even though microscopic evaluation of water samples can be tedious and time-consuming (especially for large sample sets), and requires a certain level of taxonomic expertise, this approach is employed routinely by some monitoring programs as a rapid screen for increasing levels of potentially harmful species (e.g., K. brevis; Steidinger, 1993). Automation of species identification based on morphological and other optical properties (e.g., fluorescence, light scatter, pigments) (Lorenz et al., 1999; Millie et al., 1997) is now being addressed by combining advanced imaging/ image analysis technology with flow cytometry to yield instruments capable of continuous monitoring for species of interest and their concentrations (e.g., Fluid Imaging Technologies, r FlowCAM
, 2003). Nevertheless, one of the greatest challenges in HAB species detection is the fact that many harmful/toxic species are morphologically indistinguishable from benign, closely related taxa, a critical issue from a management perspective. The need to move beyond

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identification employing only cell morphology has led to the development of molecular-based detection methods that target specific molecules such as cell surface moieties (e.g., proteins, carbohydrates, etc.) and various components of an organism’s genome. There are several approaches currently being used to frequently detect these speciesspecific classes of molecules, incorporating antibody, lectin, or oligonucleotide probes. Detection methods employing antibodies and lectins are generally (yet not exclusively) aimed at recognizing algal cell surface molecules, with positive reactions generated via a fluorescent signal from a fluorophore directly coupled to the antibody or lectin (see review by Vrieling & Anderson, 1996). Both antibodies and lectins have been applied to the labeling and detection of multiple harmful algal species (see reviews by Scholin et al., 2003, 2004; Sellner et al., 2003). It is critical to examine experimentally the potential fluctuations in target abundance as a function of the algal cells’ physiological status, since signal strength per cell and thus detection limit will vary accordingly. The two most often used strategies for visualizing these fluorescently labeled, intact cells are epifluorescence microscopy and flow cytometry. The former yields confirmatory morphological data, while the latter provides more options for automated detection (see review by Peperzak et al., 2000). Reports of incorporating antibodyor lectin-based detection into routine HAB monitoring are very limited and region-specific (e.g., Cho et al., 2000). Recently, a high throughput (96-well plate format), enzyme-linked immunosorbant assay (ELISA) using a cell-surface, species-specific antibody (Caron et al., unpublished data) has been employed to monitor Aureococcus anophagefferens, the brown tide organism. Of the three molecular level probes mentioned above, oligonucleotide probes targeting taxon-specific nucleic acid sequences have been more widely accepted by regulatory agencies (see reviews by Scholin et al., 2003, 2004; Sellner et al., 2003). These probes are most commonly designed to recognize rRNA or rDNA signature sequences in cells of a given harmful algal species and a variety of reporting systems are available for detection of positive signals. The so-called whole cell (WC) hybridization approach involves penetration of the probe into intact cells, binding to its target on the rRNA molecules, followed by visualization using a fluorescent reporter either attached directly to the probe or applied during a secondary labeling step. As with antibody- or lectin-labeled cells, epifluorescence microscopy or flow cytometry can be used to detect oligonucleotide-labeled cells. The WC technique is currently being used to detect the harmful species Alexandrium spp. and Pseudo-nitzschia spp. in New Zealand, and probes for additional species (e.g., Karenia spp., Heterosigma spp., and Fibrocapsa spp.) are also being tested for inclusion in that country’s phytoplankton monitoring programs (Cawthron Institute, 1999; Rhodes et al., 2001a, b). In addition to the WC approach a second method, referred to as “sandwich hybridization” (SH), has been employed to detect harmful algal species using gene probes. This technique involves lysis of the algal cells to release the rRNA target molecules, which are then acquired by a species-specific “capture” probe immobilized on a solid support and visualized by a second taxon-specific “signal” probe via flexible reporting systems (e.g., color, fluorescence, and chemiluminescence). While the SH method provides no morphological data, it has been successfully automated and represents a rapid, relatively high throughput approach to detecting a range of harmful species (e.g., Pseudo-nitzschia spp., Alexandrium spp., and several raphidophytes; see Scholin et al., 2003) that will be of interest to established monitoring programs when thoroughly tested and commercially available (e.g., Rhodes et al., 2001b). Moreover, a

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SH protocol has recently been incorporated onto an in situ platform called the Environmental Sample Processor (ESP) that is deployed on a stationary mooring, conducts automated SH tests for harmful algal species, and reports data back to a land-based monitoring station (Scholin et al., 2004). In addition, other gene probe-based detection strategies employing real-time PCR (polymerase chain reaction) technology are currently being developed (Bowers et al., 2000) and are anticipated to be of interest to monitoring programs.

8.8.2. Toxins The detection of toxins produced by harmful algal species involves a number of challenges. Among these are the broad chemical and structural diversity of such compounds, large differences in intrinsic potency not only between toxin classes but also among analogs of the same toxin group, and the susceptibility of toxins to biotransformation once incorporated into a biological matrix through trophic transfer processes. Moreover, certain toxins may also be found dissolved in seawater. To address these various issues, a range of detection methods have been developed, which can be grouped in three general categories: in vivo (i.e., whole-animal) bioassays, in vitro assays, and chemical analyses. In the case of several toxin groups (e.g., saxitoxins, brevetoxins, and dinophysistoxins), the most widely accepted detection method for regulatory purposes is the mouse bioassay, which involves intraperitoneal (i.p.) injection of an extract of the organism being tested. While there are a number of intrinsic and operational shortcomings (reviewed by Fern´andez et al., 2003), this in vivo mammalian bioassay is considered to yield a reasonable estimate of human toxic potency and has proven historically to be a reliable means of enforcing regulatory limits and protecting seafood consumers from toxin exposure. Nonetheless, issues such as poor sensitivity, high inter-laboratory variability, maintaining mammalian bioassay facilities, as well as mounting pressure against the use of live animals for toxicity testing are driving efforts to develop alternative in vitro methods or chemical assays suitable for regulatory use. In vitro assays for algal toxins can be either function- or structure-based, depending on the mechanism of toxin recognition, and these exist for most of the major toxin classes (see reviews by Cembella et al., 2003; Van Dolah & Ramsdell, 2001). Functional assays rely on the detection of a toxin’s biochemical activity, while the latter employ recognition of chemical structure at the molecular level, and there are advantages and disadvantages to both approaches. Examples of functional tests include receptor binding assays (Powell & Doucette, 1999; Van Dolah et al., 1994), cell-based cytotoxicity assays (Manger et al., 1995), enzyme inhibition assays (Della Loggia et al., 1999), and reporter gene assays (Fairey et al., 1999). These methods depend on toxin binding by specific biological receptors, or interference with enzyme activity and the affinity of this interaction is proportional to the intrinsic toxic potency. As a result, the assay response provides an estimate of net toxicity for all similarly functioning analogs present in a sample, yet no information about the actual identity of the toxin congeners is obtained—only their toxic activity is known. The most widely available structure-based assays for algal toxins are those using an antibody or immunological approach (Towers & Garthwaite, 2001). In certain cases, the utility of immunoassays may be limited by their lack of cross-reactivity with toxin derivatives and/or biometabolites other than that used to produce the antibody employed in the test. However, recent advances in the preparation of antibodies that allow more control

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over specificity and the incorporation of multiple antibodies into a single assay have resulted in more reliable, robust products, which are now commercially available for several toxin classes (e.g., PSP, ASP, DSP, and CFP; see Cembella et al., 2003). At present, no in vitro assays serve as the sole regulatory method for an algal toxin; nevertheless, extensive validation studies for PSP toxin tests have been completed (Jellett Rapid Testing Limited, 2004), or are planned for the near future (saxitoxin receptor binding assay) and show promise as possible alternatives to the mouse bioassay. Analytical methods based on the physicochemical properties of algal toxins primarily involve the coupling of HPLC separations with one of several detectors, including ultraviolet (UV), fluorescence, and MS (see Botana, 2000; Cembella et al., 2003; Lewis, 2003; Long & Carmichael, 2003; Luckas et al., 2003; Quilliam, 2003a, b; Scholin et al., 2003). Detection limits of these various approaches can range across orders of magnitude and are limited to those toxins for which standard reference material is available for calibration and quantification. Analytical methods have been developed for most toxins and are widely used by the research community; however, only the HPLC-UV protocol for domoic acid in shellfish has been accepted internationally as a regulatory method. In cases where an in vivo or in vitro assay yields a positive response, it is frequently of interest to establish the presence of a given toxin or toxin analog (in addition to its toxic activity), in which case LC-tandem MS provides unambiguous confirmation based on a toxin’s fragmentation pattern in the instrument, representing a unique fingerprint for that molecule. Application of recent separation technology (i.e., hydrophilic interaction liquid ion chromatography) coupled to MS has made possible the near-simultaneous resolution of more than a dozen components representing four major algal toxin groups from a common matrix in a single chromatographic run (Quilliam, 2003a). While such methods require substantial investment in equipment as well as technical expertise of the operator, concurrent detection of multiple toxin groups in a single sample is highly desirable in situations where several classes of algal toxins may co-occur.

8.9. MANAGEMENT STRATEGIES Management options for harmful algal blooms include strategies for reducing their incidence (prevention), actions to terminate or contain existing blooms (control), and steps to reduce human health and environmental consequences (mitigation) (reviewed by Boesch et al., 1997).

8.9.1. Prevention HABs associated historically with human illnesses and marine animal mortality events, such as Alexandrium blooms in the Gulf of Maine or K. brevis blooms in the Gulf of Mexico, are natural phenomena that occur on a near annual basis in the absence of direct anthropogenic influences. However, increases of nutrient loads in coastal waters over the last half-century correlate, in many instances, with increases in the occurrence of other HAB species (reviewed in Anderson et al., 2002; Sellner et al., 2003; see page 191). A key HAB prevention strategy rests fundamentally on the reduction of nutrient inputs. Numerous examples now exist to demonstrate that implementation of nutrient controls results

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in decreased phytoplankton biomass overall and a decrease in the incidence of HAB species in particular (Anderson et al., 2002). An often-cited example is the correlation between the eightfold incidence in frequency of algal blooms in Tolo Harbor, Hong Kong from 1976 to 1986, and a sixfold increase in population during that period, which was accompanied by a twoand-a-half-times increase in nutrient loading (Lam & Ho, 1989). Following the introduction of nutrient controls in the mid-1980s, the incidence of red tides in Tolo Harbor decreased significantly and the phytoplankton community shifted to one dominated by diatoms over dinoflagellates (Yung et al., 1997). The inadvertent transfer of nonindigenous HAB species and their cysts via ballast water, or in association with fish and shellfish transplanted for aquaculture purposes, are two major, preventable mechanisms by which HABs can be introduced to novel locations. Approximately ten billion tons of ballast water is transported each year, carrying with it pathogenic bacteria, invertebrates, and fish to new locations (Carlton & Geller, 1993). A number of these introductions have had significant ecological and economic impacts (Ruiz et al., 1997) or human health effects (Colwell, 1996). Ballast water transferred from Japan is believed to be the source of saxitoxin-producing dinoflagellates introduced in Australia (Hallegraeff, 1998) (page 192) and has been demonstrated to transport viable diatoms and dinoflagellates from Mexico to Hong Kong (Zhang & Dickman, 1999). Consequently, procedures for exchanging ballast waters at sea and/or shipboard treatment to remove ballast-associated biota have been implemented, although development of improved techniques remains a priority (National Research Council, 1996). When shellfish are transported from a toxic area to clean waters, they may release cysts and/or motile algal cells which may in turn sow a bloom of the toxic algae if environmental conditions in the receiving waters are conducive to their growth (Scarratt et al., 1993). Such introductions have been documented in the port of Sete, in the south of France, where productive shellfish growing waters are now plagued with annual PSP and DSP producing blooms. Many countries have established regulations to prohibit relocation of mussels from potential PSP risk areas to other areas in an effort to control spreading of blooms (Shumway, 1990). However, with the vast increases in aquaculture activities worldwide, diligent attention to these issues will be required.

8.9.2. Control Strategies to eliminate HABs are somewhat controversial due to their inherent potential to adversely affect other components of the ecosystem. Possible approaches include chemical treatment, physical disruption of blooms, and biological control. The use of chemical controls, although employed to control algal growth in freshwater ponds, has not been successfully employed in estuarine and marine situations and is not currently viewed as a suitable strategy because of the likelihood of nonspecific toxicity toward other organisms. Early attempts at controlling the Florida red tide with the application of copper sulfate highlight this problem (Rounsefell & Evans, 1958). Currently employed control strategies consist primarily of physical procedures to disrupt blooms, particularly with the addition of flocculants to physically remove HAB cells from the water column. Clays have been employed for HABs on a massive scale in fish farming areas of the Korean peninsula (Na et al., 1996). A variety of clays and chemical flocculants have been shown to successfully remove both K. brevis and brevetoxins from the water column (Pierce et al., 2004). However, potentially adverse effects of flocculants on benthic organisms, either directly (Archambault et al., 2002) or as

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the result of toxin transfer to the benthic community (Lewis et al., 2003), must now be tested. Both brevetoxins and domoic acid are known to enter the benthic food web following the demise of blooms, and such benthic routes can impact even top trophic level predators such as sea otters (Lefebvre et al., 2002) and manatees (Flewelling et al., 2003). Another proposed approach to physically disrupting blooms is to alter water conditions where a bloom may be developing. For example, it has been proposed that K. brevis red tides in the area of the Caloosahatchee River outfall in southwest Florida, USA might be disrupted by increasing freshwater outflow through dams and thus decreasing salinity in the embayments to levels below the tolerance limit for the dinoflagellate. This strategy, although considered by the South Florida Water Management District, has not yet been attempted as a control strategy (Landsberg, personal communication). Lastly, the use of biological controls such as grazers or algicidal viruses, bacteria, or parasites (Doucette et al., 1999; Nagasaki et al., 1999) remains experimental.

8.9.3. Mitigation Mitigation consists of procedures that can be followed to reduce the health risk and economic losses from HABs that are not successfully prevented or controlled. The early detection (see previous section) of harmful algal species and toxins is a critical component of mitigation of HAB effects. Many coastal states have in place algal monitoring programs designed to anticipate and give prior notice of impending toxic blooms. However, the existing monitoring programs depend upon the time-consuming task of microscopic examination and enumeration of algal cells present in a water sample, followed by toxin testing in shellfish when a prescribed cell concentration has been detected. Therefore, significant research efforts are currently targeted at the development of remote-detection methods, capable of real-time monitoring and suitable for deployment on moored buoys or robotic submersible vehicles. Several different approaches have been taken, including the use of species-specific DNA probes, species- and toxin-specific antibody probes, in situ mass spectrometry, and in situ videoimaging systems that can be deployed on moored buoys or unmanned mobile submersibles (reviewed by Scholin et al., 2004; Sellner et al., 2003). In addition, satellite imaging of ocean color and spectral analysis of pigment profiles have recently been successfully employed to identify offshore blooms of K. brevis in the Gulf of Mexico prior to their movement into coastal waters. Coupled with predicted wind fields, this information has enabled the accurate forecasting of landfall of blooms, giving coastal managers early warning of toxic conditions (Stumpf et al., 2003).

8.10. FUTURE With increasing anthropogenic pressures upon our coastal habitats and oceanic resources, the threat of HABs to public and natural resources health is an expanding problem. Technological advances in monitoring and detection have expanded our capabilities to predict HAB events, but their management and mitigation is not an easy task. Combined multidisciplinary efforts, risk assessments, and impact strategies need to be developed and implemented at multiple scales to prevent further economic and ecological losses. Anthropogenic changes will continue to enhance the distribution of these blooms and will allow for the appearance of

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new harmful and toxic species. It is time to acknowledge the extreme range of their effects on human and natural resources and to begin to minimize the severity of their ecological and economical impacts.

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9 Pathogenic Vibrio Species in the Marine and Estuarine Environment Carla Pruzzo, Anwar Huq, Rita R. Colwell, and Gianfranco Donelli 9.1. INTRODUCTION The genus Vibrio includes more than 30 species, at least 12 of which are pathogenic to humans and/or have been associated with foodborne diseases (Chakraborty et al., 1997). Among these species, Vibrio cholerae, serogroups O1 and O139, are the most important, since they are associated with epidemic and pandemic diarrhea outbreaks in many parts of the world (Centers for Disease Control and Prevention, 1995; Kaper et al., 1995). However, other species of vibrios capable of causing diarrheal disease in humans have received greater attention in the last decade. These include Vibrio parahaemolyticus, a leading cause of foodborne disease outbreaks in Japan and Korea (Lee et al., 2001), Vibrio vulnificus, Vibrio alginolyticus, Vibrio damsela, Vibrio fluvialis, Vibrio furnissii, Vibrio hollisae, Vibrio metschnikovii, and Vibrio mimicus (Altekruse et al., 2000; Høi et al., 1997). In the USA, Vibrio species have been estimated to be the cause of about 8000 illnesses annually (Mead et al., 1999). Vibrios can be classified as either halophilic or nonhalophilic, depending on their requirement for NaCl for optimal growth (Thompson et al., 2004). They are free-living bacteria in the aquatic environment throughout the world. They tend to be more common in warmer waters, notably when temperatures rise above 17 ◦ C and, depending on the species, they tolerate a wide range of salinities (Wright et al., 1996). Given their abundance in water, Vibrio species are also commonly isolated from fish and shellfish, with 100-fold higher concentration in filter-feeding shellfish, such as oysters, than in the surrounding water (Wright et al., 1996). During the warm summer months, virtually 100% of oysters can carry V. vulnificus and/or V. parahaemolyticus (Cook et al., 2002b; Motes et al., 1998; Wright et al., 1996). “Epidemic” strains of V. cholerae, which carry specific virulence genes, cause the disease “cholera,” while “nonepidemic” strains are mainly associated with septicemia, gastroenteritis, and wound infections (Levine & Griffin, 1993). However, morbidity due to nonepidemic strains is relatively low, and, in some instances, it is unclear whether isolation of the organism Carla Pruzzo • Department of Biology, University of Genova, Italy. Anwar Huq, Rita R. Colwell • Center of Marine Biotechnology, University of Maryland Biotechnology Institute, Baltimore, MD, USA and UMIACS, University of Maryland, College Park, MD, USA. Gianfranco Donelli • Department of Technologies and Health, Istituto Superiore di Sanit`a, Rome , Italy.

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represents asymptomatic colonization or infection. In fact, given the ubiquitousness of the genus in the environment, asymptomatic colonization may occur.

9.2. SPECIES PATHOGENIC FOR HUMANS IN SEAWATER 9.2.1. Isolation and Identification, Significant Virulence Properties, and Transmission Routes V. cholerae is the most important pathogenic species, because of the production of a potent enterotoxin, the cholera toxin (CT), that disrupts the ion transport of intestinal epithelial cells. The subsequent loss of water and electrolytes leads to severe diarrhea and vomiting, distinctive characteristics of cholera, and results in severe dehydration. Other vibrios, reviewed below, may produce virulence factors or possess pathogenicity factors that are yet not discovered; thus, from a public health perspective, such species are precautionally considered to be potential pathogens (Barbieri et al., 1999; Chakraborty et al., 1997). 9.2.1.1. Vibrio cholerae: O1 and O139 Serogroups Although V. cholerae comprises more than 200 O-antigen-based serogroups (Shimada et al., 1994), only the O1 and relatively recently recognized O139 serogroup are known to cause the massive epidemics of cholera. The O1 serogroup can be further divided into two biotypes, called O1 “El Tor” and O1 “classical,” determined by physiological properties, such as polymyxin B resistance, number of CT-encoding genes, hemolysin activity, and the presence of the mannose-sensitive hemagglutinin (MSHA); each biotype may have three different serotypes termed “Inaba,” “Ogawa,” and “Hikojima” (Kaper et al., 1995). Seven distinct pandemics of cholera have occurred since the onset of the first pandemic in 1817 (Barua, 1992). While the early pandemics have never been associated with a specific strain variant, the fifth and sixth pandemics have been shown to have been caused by V. cholerae serogroup O1, of the biotype “classical,” while the seventh pandemic was caused by serogroup O1 biotype “El Tor.” In 1992, a new serogroup, O139, was recognized to be the cause of a large epidemic, first in Madras, India (Ramamurthy et al., 1993) and thereafter in Bangladesh (Albert et al., 1993) displacing the well known strains of O1 serogroup. This epidemic has been designated as the eighth pandemic (Faruque et al., 1998). However, the O1 strains reemerged in 1994 and thereafter there initiated a series of disappearances and reemergences of the two serogroups (Faruque et al., 2003). In important virulence characteristics, specifically, the CT and toxin-coregulated pilus (TCP) (see below) of V. cholerae O139 are indistinguishable from those of the typical El Tor V. cholerae O1 strains. Interestingly, V. cholerae O139 does not produce O1 LPS and produces a polysaccharide capsule which may contribute to the case of septicemia as having been associated with this serogroup (Jesudason et al., 1993). Cholera is the result of a multifactorial process, involving the coordinated expression of colonization factors and toxins. Understanding the role played by each of the virulence factors is facilitated by the recently published complete genome sequence of the two chromosomes of V. cholerae (Heidelberg et al., 2000). The larger chromosome includes genes encoding for the two major virulence factors, CT and TCP. In turn, CT is encoded by genes located within a lysogenizing filamentous bacteriophage (CTx) and TCP, an essential colonization factor

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encoded by tcpA, is also the receptor for CTx (Waldor & Mekalanos, 1996). In addition to TCP, several V. cholerae gene products have been shown to be important for colonization of the small intestine, primarily in the infant mouse and the adult rabbit model of cholera. These include accessory colonization factors (ACFs) (Peterson & Mekalanos, 1988), mannose– fucose-resistant hemagglutinin (Franzon et al., 1993), TolC (Bina & Mekalanos, 2001), outer membrane porins (Provenzano et al., 2001), the O-antigen of the lipopolysaccharide (LPS) (Baselski et al., 1979), and attributes of the LPS core region (Nesper et al., 2001). Extensive study of the molecular basis of cholera pathogenesis has revealed that expression of the most important virulence factors, namely CT and TCP, is coordinated by a unique regulatory system. In this regulatory cascade, the inner membrane DNA-binding proteins, ToxR and TcpP, control the expression of ToxT, a transcriptional activator that specifically activates the ctx and tcp gene clusters, as well as additional genes (Dromigni et al., 2002; Kondo et al.,1994). Expression of the two main virulence factors by the ToxR regulon is jointly regulated and deeply influenced by environmental and physiological signals, including temperature, osmolarity, pH, amino acids, and bile. In fact, these parameters are known to modulate the expression of the ToxR regulon, acting at different levels of the regulatory cascade (Skorupski & Taylor, 1997). CT is a potent enterotoxin required for the induction of secretory diarrhea. The structural genes for CT are encoded within a lysogenic phage (CTx) integrated into the large chromosome (Waldor & Mekalanos, 1996). TCP is a type IV pilus required for intestinal colonization. The structural genes for TCP are encoded within a pathogenicity island that displays some features of a transmissible genetic element (Kaysner et al., 1987). The CTx can transfer horizontally to non-toxin-producing strain as long as the recipient expresses TCP, which acts as the phage receptor. Two inner membrane proteins, ToxR and TcpP, interact with the cytosolic protein ToxT to coordinately regulate the expression of TCP and CT. During infection, ToxR and TcpP induce expression of CT and TCP after sensing the as yet unknown factors in the intestinal microenvironment. In fact, V. cholerae has been demonstrated to interact with the gut microenvironment, using a novel transcriptional reporter system to track the in vivo expression of CT and TCP, observed during infection of suckling mice (Lee et al., 1999). After entry into the intestinal lumen, the vibrios sense a luminal factor(s) that, inducing low-level expression of TCP, enables the organisms to adhere to the intestinal mucosa (presumably to the glycocalyx) and then to sense the subsequent environmental signal. This second signal depends on microbial adherence to the mucosa and induces increased levels of TCP expression, enhanced intestinal colonization, and the production and secretion of the enterotoxin CT. Both ToxR and TcpP are critical for these events. However, the identity of factors sequentially activating ToxR and TcpP during in vivo infection and their origin (resident microbial flora, or intestinal mucosa of the host) are still unknown. To induce disease, cholera toxin, after its release into the gut lumen, must enter the intestinal epithelial cells at the apical membrane level and activate adenylyl cyclase at the cytoplasmic surface of the basolateral membrane. The massive intestinal salt and water secretory response characteristic of cholera is presumably amplified by interaction of the toxin with the enteric nerves and subsequent local release of neurotransmitters. Molecular determinants driving toxin entry into host intestinal epithelial cells are encoded entirely within the structure of the protein toxin itself. In fact, V. cholerae itself does not enter the intestinal mucosa (Mekalanos & Sadoff, 1994) even if there is a possible active role of the microorganism in assisting toxin

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delivery into the cell cytoplasm through a mechanism, such as type III secretion, as has been hypothesized. The crystal structures of CT and the closely related Escherichia coli heat-labile enterotoxin ˚ resolution (Sixma et al., 1993). Both toxins belong to the AB5 LTI have been solved at 2.2-A subunit family that also includes the pertussis toxin and E. coli Shiga-like verocytotoxins, the latter responsible for hemolytic uremic syndrome. In cholera toxin, the A subunit is composed of two major structural peptidic domains, A1 and A2 . These peptides are linked by an exposed loop containing a serine protease-sensitive site and a single disulfide bond. The A2 peptide (∼5 kDa) binds the A1 peptide to the B subunit and contains a COOH-terminal KDEL motif (RDEL in LTI) that protrudes below the pentameric B subunit on the side that binds GM1 at the cell surface (Sixma et al., 1993). The KDEL motif is recognized as a sorting signal that allows the efficient retrieval of endogenous luminal endoplasmic reticulum (ER) proteins from postER compartments of the eukaryotic cell. The A1 peptide (∼22 kDa) is the enzymatic subunit that must dissociate from the B subunit, translocate across the cellular membrane, and activate inside the cell adenylyl cyclase by catalyzing the ADP-ribosylation of GTPase Gsα . The B subunit (∼55 kDa) is a highly stable pentameric ring composed of five identical peptides (∼11 kDa). This subunit is responsible for the toxin interaction with the host cell plasma membrane exhibiting a specific and high-affinity binding to the oligosaccharide domain of ganglioside GM1 . This binding is stoichiometric, with one pentameric B subunit cross-linking five GM1 gangliosides at the cell surface. Since CT is not a pore-forming toxin, to enter intestinal cells it exploits the membrane machinery endogenous to the host epithelial cells through a complex pathway involving apical endocytosis and retrograde membrane traffic through Golgi cisternae to ER (Lencer et al., 1995a; Majoul et al., 1996; Orlandi, 1997). To allow the A1 peptide to unfold and translocate into the cytosol, CT must, in fact, enter the ER. Thus, after membrane translocation, the A1 peptide may move by diffusion through the cytosol to the adenylyl cyclase complex on the cytoplasmic surface of the basolateral membrane if the A1 peptide breaks away from the membrane after translocation, or by vesicular transport back out of the secretory pathway if the A1 peptide remains membrane associated (Lencer, et al., 1999). Different from the A1 peptide, B subunit does not translocate across the cell membrane, remaining presumably bound to GM1 and eventually moving back out to the cell surface by vesicular traffic (Lencer et al., 1995b). Thus, the B subunit moves from its initial cell binding site on the apical surface to the basolateral surface by translocating through the Golgi cisternae and possibly reaching the ER; this process has been termed “indirect” transcytosis (Lencer et al., 1995b). Other important virulence factors include the Vac-like toxin, an accessory cholera enterotoxin (Ace), zonula occludens toxin (ZOT), and the secreted proteins RTX toxin and HA protease. The phenomenon of vacuolation of mammalian cells that occurs in response to bacterial secreted proteins was first described for Helicobacter pylori in the action of the secreted cytotoxin VacA (Cover & Blaser, 1992; Leunk et al., 1988). The aerolysin of Aeromonas hydrophila and the hemolysin of Serratia marcescens have also been reported as cell vacuolating agents (Abrami et al., 1998; Hertle et al., 1999). With respect to V. cholerae, a cytotoxin designated VcVac was found to cause vacuolation in Vero cells. It was originally detected in the pathogenic O1 Amazonia variant of V. cholerae and later shown to be present in environmental strains and a few El Tor strains. Results of a comparison of VcVac production by various strains suggested that hemolysin was responsible for the vacuolating phenotype. Genetic experiments subsequently established a firm correlation between vacuolation and hemolysin production.

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The mammalian cell vacuolating activity of the V. cholerae hemolysin is a new property of this protein, representing a previously unknown type of interaction between V. cholerae and its host (Coelho et al., 2000). Accessory cholera enterotoxin (Ace) was proposed as a putative toxin of V. cholerae by Trucksis et al. (1993). Preliminary studies using crude toxin extracts in animal models indicated that Ace increases transcellular ion transport, which is supposed to play a role in causing diarrhea. The Ace toxin is harbored in the V. cholerae chromosomal core region, encoded by the filamentous bacteriophage CTx (Waldor & Mekalanos, 1996), and induces chloride and bicarbonate secretion from gut epithelial cells and causes fluid secretion in ligated rabbit ileal loops, suggesting its role as a pore-forming toxin (Trucksis et al., 1997). In vitro studies provided evidence for Ace being a Ca2+ -dependent agonist active against the apical membrane of polarized intestinal epithelial cells T84 by stimulating anion secretion. These data are consistent with the above mentioned reports on its secretory activity in animal models (Trucksis et al., 2000). Zot is a single polypeptide chain of 44.8 kDa encoded by the filamentous bacteriophage CTx present in toxigenic strains of V. cholerae (Baudry et al., 1992; Waldor & Mekalanos, 1996). Zot increases the permeability of the small intestine by affecting the structure of epithelial tight junctions, providing a paracellular route for the traffic of macromolecules (Fasano et al., 1995). This effect is dependent on the binding of Zot to specific receptors, as was clearly demonstrated by in vivo models and in vitro experiments on cultured epithelial cell lines (Fasano et al., 1991; Uzzau et al., 2001). Zot does not cause tissue damage and its timeand dose-dependent effect on intestinal permeability is reversible (Fasano et al., 1995; Felter et al., 1969). The Zot mode of action on tight junctions involves a rearrangement of the cytoskeleton due to protein kinase C alpha-dependent F-actin polymerization (Fasano et al., 1995). A mammalian analog of Zot, named zonulin, has been found on intestinal epithelial cells and reported to play a role as endogenous regulator of tight junctions (Fasano, 2001; Wang et al., 2000). Experimental stimuli, such as bacterial loading or gut tissue injury, have been shown to cause zonulin release in the intestinal lumen, which then binds to its receptor and modulates tight junctions (Fasano & Uzzau, 1997). Since Zot and zonulin bind to the same receptor on intestinal epithelial cells and share a common binding motif at their N termini (DiRita, 1992; Wang et al., 2000), Zot results in mimicking a physiological molecule, i.e. zonulin, in the regulation of tight junctions. A novel V. cholerae toxin that belongs to the RTX (repeat in toxin) family of toxins generally produced by several pathogenic Gram-negative bacteria was recently described (Lin et al., 1999). The RTX toxins are important virulence factors widely disseminated among Gramnegative bacteria (Dalsgaard et al., 2000) characterized by hemolytic, leukotoxic, and leukocyte-stimulating activities (Braun et al., 1993; Welch et al., 1995). In V. cholerae, the RTX toxin gene cluster encodes the presumptive cytotoxin RtxA, the acyltransferase RtxC, and an associated ATP-binding cassette constituted by the two transporter proteins RtxB and RtxD. Although its activity is independent, this system for toxin transport is linked to the core element in the V. cholerae genome (Lin et al., 1999). To overcome the protective mucus barrier covering the gastrointestinal epithelium, V. cholerae produces a soluble Zn2+ -dependent metalloprotease (mucinase), the hemagglutinin/protease encoded by hapA (Booth & Finkelstein, 1986). Hap is secreted through the general secretory pathway (Sandkvist et al., 1997). DNA sequence homology revealed that

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Hap is closely related to the Pseudomonas aeruginosa elastase and the metalloproteases secreted by Legionella pneumophila and V. vulnificus (Miyoshi & Shinoda, 2000). The size of hapA ORF compared with the molecular mass of secreted Hap suggests that its release involves sequential cleavage of a signal peptide and a propeptide (Heidelberg et al., 2000). Poor protease activity can be revealed in mutant Vibrio strains which lack a functional hapA gene, suggesting that this gene encodes the most important protease secreted by V. cholerae. Hap can proteolytically activate the A subunit of cholera toxin (Booth et al., 1984), the El Tor cytolysin/hemolysin (Nagamune et al., 1996), and hydrolyze several important proteins such as fibronectin, lactoferrin and mucin (Fouz et al., 1992). It has also been reported that Hap is able to cause morphological and functional alterations in the paracellular barrier of cultured MDCK-1 and T84 intestinal epithelial cells (Mel et al., 2000; Wu, et al., 1996) by acting on tight-junction-associated proteins (Wu et al., 2000). 9.2.1.2. Vibrio cholerae: Non-O1/non-O139 Serogroups Non-O1/non-O139 V. cholerae strains are occasionally isolated from cases of diarrhea that are mild to moderate in severity and from a variety of extraintestinal sites, such as wounds, ear, respiratory tract, and cerebrospinal fluid (Morris Jr., 1990; Kaper et al., 1995). The majority of these strains do not produce CT and are not associated with epidemic diarrhea; however, they produce a number of putative virulence factors, including Shiga-like cytotoxin (O’Brian et al., 1984), a heat-stable enterotoxin (Arita et al., 1986) and hemolysins (Yamamoto et al., 1986). These organisms are considered to be autochthonous components of the bacterial flora of estuarine waters. Isolates of V. cholerae non-O1/non-O139 are also capable of surviving and multiplying in a wide range of seafoods (Shimada et al., 1993). This characteristic may explain why most human infections caused by V. cholerae non-O1 are associated with exposure to aquatic environments or to recent consumption of seafood. Recent studies have shown that virulence genes, or their homologues, are dispersed among environmental strains of V. cholerae belonging to diverse serogroups, which appear to constitute an environmental reservoir of virulence genes (Chakraborty et al., 2000). However, in some cases, no link with the aquatic environments could be established, indicating the existence of other sources of infection. In view of reports of isolation of V. cholerae non-O1 from food-production animals in the Netherlands (Shimada et al., 1993), such as calves and lambs with diarrheal disease, a role of these animals as a reservoir for V. cholerae non-O1 would be unusual but should be assessed. 9.2.1.3. Vibrio parahaemolyticus V. parahaemolyticus is a halophilic bacterium that naturally inhabits marine and estuarine waters and was first identified as a cause of foodborne illness in Japan in 1950. This common agent of gastroenteritis, which is still the leading source of foodborne illness in Japan and throughout Asia (Lee et al. 2001; Wong et al., 2000), can also cause wound infections and septicemia in susceptible hosts (Morris & Black, 1985). Although V. parahaemolyticus is now recognized as one of the most important agents of seafood associated gastroenteritis, not all strains of this species are considered truly pathogenic (Mead et al., 1999). V. parahaemolyticus outbreaks in the United States have been associated with consumption of raw or undercooked shellfish (Bean et al., 1998). The main virulence factor that, for a long time was epidemiologically associated with disease, is one of the hemolysins produced by V. parahaemolyticus.

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Almost all V. parahaemolyticus strains isolated from clinical specimens exhibit, when streaked on a special blood agar medium (Wagatsuma agar), a beta-type hemolysis (Miyamoto et al., 1969), and this characteristic has been named the Kanagawa phenomenon (KP) (Miyamoto et al., 1969). This hemolysin was subsequently defined as a thermostable direct hemolysin (TDH), since it is not inactivated by heating at 100 ◦ C for 10 min and its hemolytic activity is not enhanced by addition of lecithin, indicating direct action on erythrocytes (Sakurai et al., 1973). Biological activities of the TDH include hemolysis of erythrocytes of several animal species, cytotoxicity, lethality for small experimental animals, and increased vascular permeability when tested on rabbit skin (Honda, & Miwatani, 1988; Takeda, 1983). TDH is a proven virulence factor (Nishibuchi et al., 1992), which occurs in over 90% of the clinical strains isolated both in the United States (Okuda et al., 1997) and worldwide (Miyamoto et al., 1969; Nishibuchi et al., 1985; Wong et al., 2000). In vitro studies on TDH cytotoxicity and enterotoxicity demonstrated how the toxin modulates cytoskeletal organization and calcium homeostasis (Fabbri et al., 1999; Raimondi et al., 2000) while recent data, obtained by Baffone et al. (2005) in the rat ileal loop model, reported a pivotal role of calcium ions in V. parahaemolyticus TDH+ -induced secretion. The tdh gene, usually but not exclusively located in the chromosome, was first cloned from the KP-positive strain WP1. All of the cloned tdh genes encode predicted protein products (including signal peptides) composed of 189 amino acid residues, which exhibit hemolytic and other biological activities. The nucleotide sequences of the various tdh genes are well conserved (>97% identity), and the protein products are immunologically indistinguishable. In an outbreak of gastroenteritis in the Maldives, KP-negative isolates of V. parahaemolyticus were found to produce a TDH-related hemolysin (TRH) but not TDH (Hoshino et al., 1998). Biological, immunological, and physicochemical characteristics of TRH are similar but not identical to those of TDH (Hoshino et al., 1998). Most clinical isolates from the U.S. Pacific Coast have been reported to possess both tdh and trh (Okuda et al., 1997). Environmental isolates possessing trh have been reported in Asia (Wong et al., 2000), only recently in the U.S. Gulf Coast, but not yet in the Atlantic Coast (Cook et al., 2002b; Thompson & Vanderzant, 1976). Strains possessing trh almost always produce urease, which is uncommon for other V. parahaemolyticus strains (Okuda et al., 1997) and these genes may be part of a pathogenicity island (Park et al., 2002). Neither urease nor trh has been shown to enhance the virulence of those strains that also possess tdh, but urease has been found to increase acid tolerance in Yersinia enterocolitica (DePaola et al., 2000) and Helicobacter pylori (Turbett et al., 1992). The recently sequenced genome of a clinical V. parahaemolyticus strain RIMD2210633 consists of a large chromosome (chr1) and a small chromosome (chr2) (Makino et al., 2003). The former carries all the essential cell functions and housekeeping genes, whereas the latter carries genes that might be involved in overall fitness of the bacterium or in conferring advantage(s) in diverse environments. For example, V. parahaemolyticus has two types of flagella, polar and lateral. The constitutively expressed polar flagella are on chr1, whereas genes encoding for lateral flagella, required for swarming on solid surfaces, are on chr2. Comparison of the V. parahaemolyticus genome with that of V. cholerae showed many rearrangements within and between the two chromosomes; moreover the small chromosomes were found to carry a higher proportion of genes unique to each Vibrio species. Genes for the type III secretion system (TTSS) were identified in the genome of V. parahaemolyticus; V. cholerae does not have these genes. The TTSS is a central virulence factor of diarrhea-causing bacteria, such as Shigella, Salmonella, and enteropathogenic E. coli, which cause gastroenteritis by invading or intimately

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interacting with intestinal epithelial cells. It is concluded that V. parahaemolyticus and V. cholerae use distinct mechanisms to establish infection (Makino et al., 2003). This finding explains clinical features of V. parahaemolyticus infections, which commonly include inflammatory diarrhea and, in some cases, systemic manifestations such as septicemia, distinct from those of V. cholerae infections, which are generally associated with noninflammatory diarrhea. The Kanagawa test, used in early studies to identify strains producing TDH (Miyamoto et al., 1969), has recently been replaced by the more efficient diagnostic methods, such as PCR (Bej et al., 1999; Tada et al., 1992) and DNA probe assays (McCarthy et al., 2000; Wong et al., 2000) that specifically target the gene encoding TDH (tdh). Recently, these and other genotypic assays have been applied to environmental studies (Cook et al., 2002a; DePaola et al., 1994) and seafood surveys (Coote, 1992; Falbo et al., 1999). 9.2.1.4. Vibrio vulnificus V. vulnificus is found in estuarine environments and causes primary septicemia and wound infections in humans. Primary septicemia usually occurs through ingestion of raw shellfish, especially oysters, by persons who are predisposed to infection by increased serum iron levels or who are immunocompromised. Wound infections, however, have been shown to occur in otherwise healthy individuals who come in contact with this bacterium via contamination of a previously inflicted wound or one incurred in an estuarine environment. In the United States, V. vulnificus is the leading cause of death associated with consumption of seafood, resulting from its ability to cause severe wound infections and sepsis in patients who are immunocompromised or have underlying liver disease (Blake et al., 1979; Shapiro et al., 1998). Several putative virulence factors, such as the cytolysin-hemolysin, lipopolysaccharide, capsule, and siderophores, have been identified in V. vulnificus (Linkhous & Oliver, 1999; Simpson et al., 1987). Although the frequencies of occurrence of these factors are similar among clinical and environmental isolates, most strains examined have shown different genotypes. Unfortunately, their DNA fingerprints have not been useful in virulence prediction (Buchrieser et al., 1995; Tamplin et al., 1982; Warner & Oliver, 1999). Several studies have reported the presence of V. vulnificus, as well as other estuarine vibrios, in different coastal waters (Chan et al., 1986; DePaola et al., 2003; Keasler & Hall, 1993; Kennedy et al., 1970; Levine & Griffin, 1993; Oliver et al., 1982, 1983; Tamplin et al., 1982). 9.2.1.5. Vibrio alginolyticus V. alginolyticus is a ubiquitous organism isolated in seawater and from different marine organisms as part of their saprophytic microbiota (Carli et al., 1993). This species has also been recognized for many years as a pathogen of both humans and marine animals (Blake et al., 1980; Cook et al., 2002a). The disease caused by this microorganism mainly affects individuals who are in direct contact with seawater or handle shellfish, and its incidence significantly increases during warmer months (Morris & Black, 1985). V. alginolyticus is commonly associated with ear infections (otitis media and otitis externa) and superficial wounds resulting from prior exposure to contaminated water sources (Pezzlo et al., 1979). This microorganism has rarely been associated with respiratory tract infections or bacteremia, which only in immunocompromised subjects may lead to death (Jean-Jacques et al., 1981).

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9.2.1.6. Vibrio damsela V. damsela is a vibrio species isolated from damselfish (Chromis punctipinnis) and first described by Love et al., (1981). It has been described as a human pathogen (Clarridge & Zighelboim-Daum, 1995; Morris Jr et al., 1982) that may cause fatal infections, and its pathogenicity has been partly attributed to the production of a powerful cytolysin (Lee et al., 1981). The organism has been associated with disease in a number of marine animals, including the bottlenose dolphin (Tursiops truncatus) (Gamal`eia, 1888), leatherback turtles (Dermochelys coriacea) (Obendorf et al., 1987), damselfish (C. punctipinnis) (Love et al., 1981), yellowtail (Seriola quinqueradiata) (Sakata et al., 1989), sea bream (Sparus aurata) (Vera et al., 1991), barramundi (Lates calcarifer) (Renault et al., 1994), brown shark (Carcharhinus plumbeus) (Grimes et al., 1985), and, more recently, farmed turbot (Scophthalmus maximus) (Fujioka et al., 1988). In addition, V. damsela has been shown to be a member of the microflora of healthy carcharhinid sharks. 9.2.1.7. Vibrio fluvialis V. fluvialis is a halophilic bacterial pathogen isolated from river and estuarine water, mollusks, crustaceans, and fishes and from humans with diarrhea. In fact, outbreaks and sporadic cases of diarrhea and gastrointestinal illness, usually associated with consumption of raw or improperly cooked seafood, have been frequently reported. In the United States, recent data show that this species was responsible for 82 of the 1584 Vibrio infections reported to the Centers for Disease Control and Prevention (http://www.cdc.gov/) from 1997 to 2000. Since clinical symptoms of gastroenteritis caused by V. fluvialis and V. cholerae are very similar, the ability of V. fluvialis to cause diarrhea has been evaluated in the various animal models originally designed for V. cholerae. V. fluvialis produces a toxin that elongates Chinese hamster ovary (CHO) cells, a nonhemolytic CHO cell-killing cytotoxin, a protease, and a cytolysin active against rabbit erythrocytes (Lockwood et al., 1982). Lockwood et al. (1982) showed that partially purified preparations of the elongation factor, its protease, and its cytolysin/hemolysin induced fluid accumulation in suckling mice. V. fluvialis enhanced intestinal fluid accumulation in rabbit ileal loops (Rahim & Aziz, 1994) or when fed to suckling mice (Nishibuchi et al., 1983). However, it is not clear what roles the above mentioned virulence factors play in diarrheal disease, since only the enterotoxigenic El Tor-like hemolysin has been just recently purified (Kothary & Kreger, 1985). 9.2.1.8. Vibrio furnissii The species V. furnissii was described by Brenner and co-workers in 1983 (Brenner et al., 1983) to circumscribe the biogroup 2 strains of V. fluvialis that produced gas during carbohydrate fermentation (Lee et al., 1978). DNA relatedness studies also supported a recognition of V. furnissii as a species separate from V. fluvialis (Brenner et al., 1983). V. fluvialis has been associated with sporadic cases of gastroenteritis (Klose, 2001; Lee et al., 1978; Levine & Griffin, 1993), and suggested as a cause of outbreaks of diarrhea (Islam et al., 1994). Diarrhea caused by V. fluvialis has been associated with the consumption of seafood, especially raw shellfish (Levine & Griffin, 1993). However, the role of V. furnissii as an enteric pathogen remains

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to be established, since this microorganism has frequently been isolated from the estuarine environments, but only occasionally from human cases of diarrhea (Brenner et al., 1983). 9.2.1.9. Vibrio hollisae V. hollisae is a halophilic bacterium that was first described in 1982 by Hickman et al., (1982). This bacterium produces toxins and invades host epithelial cells (Miliotis et al., 1995), causing both gastroenteritis and septicemia (Abbott & Janda, 1994). In the majority of cases reported to date, V. hollisae has been isolated from individuals with no underlying illness, but who had consumed seafood, such as raw shellfish (Abbott & Janda, 1994; Carnahan et al., 1994; Gribbon & Barer, 1995; Hlady & Klontz, 1996; Levine & Griffin, 1993; Lowry et al., 1986; Morris et al., 1982). Although the exact pathogenic mechanism of V. hollisae is not known, it has been shown to produce a hemolysin similar to that of the thermostable direct hemolysin of V. parahaemolyticus (Nishibuchi et al., 1988; Yoh et al., 1986). Compared to other species of the genus, V. hollisae is atypical because it is arginine dihydrolase-negative, lysine and ornithine decarboxylasenegative, and cannot grow on thiosulfate–citrate–bile salts–sucrose (TCBS) agar. Further, when 506 Vibrio strains were grouped by FAFLP (fluorescent amplified fragment length polymorphism analysis) to define their similarity pattern, the V. hollisae strains remained unclustered (Thompson et al., 2001). When 16S rDNA sequence analysis was done, V. hollisae strains fell between the Photobacterium species and strains of a newly described genus, Enterovibrio (Thompson et al., 2002). After investigation using both 16S rDNA sequencing and phenotypic analysis, it was concluded that V. hollisae should be placed in a new genus, for which the name Grimontia gen. nov. was proposed (Thompson et al., 2003). 9.2.1.10. Vibrio metschnikovii V. metschnikovii was first described in 1888 (Gamal`eia, 1888) and redefined as a new Vibrio species in 1978. V. metschnikovii is an oxidase negative, nitrate negative species reported to be widely distributed in the environment. Lee and Ruby (1995) isolated this bacterial species from human feces, as well as from estuaries, rivers, sewage, cockles, oysters, lobsters, and a bird dead from a cholera-like disease. Jean-Jacques et al. (1981) isolated V. metschnikovii from the blood of an elderly woman with peritonitis and an inflamed gallbladder. 9.2.1.11. Vibrio mimicus V. mimicus is an atypical strain of V. cholerae. It was regarded as non-sucrose-fermenting V. cholerae non-O1 before 1981. Thereafter, it was frequently isolated from patients and concluded to be pathogenic, causing diarrhea and abdominal cramps. V. mimicus is known to produce many toxins, such as cholera toxin, heat stable enterotoxin, phospholipase, hemolysin, and hemagglutinin (Alam et al., 1996; Nishibuchi & Seidler, 1983; Spira & Fedorka-Cray, 1984). Previously characterized hemolysin and phospholipase of V. mimicus revealed that the deduced amino acid sequences of hemolysin and phospholipase showed high similarity to those of V. cholerae (Kaper et al., 1995; Kim et al., 1997). Subsequently, rare strains of V. mimicus and V. cholerae (non-O1 strains) were isolated from diarrheal patients, but these

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strains were found to possess the gene tdh encoding the thermostable direct hemolysin of V. parahaemolyticus (Nishibuchi & Kaper, 1995).

9.2.2. Symptoms of Infection, Disease Features, and Clinical Treatment The description of cholera and its symptoms traces back thousands of years. However, in modern times, the disease and its causative agent were first described by the Italian scientist Filippo Pacini in 1864 during a severe epidemic in Florence. Today, cholera continues to be an important public health problem notably in developing and underdeveloped countries of Africa, Asia, and South America. After the ingestion of contaminated water or food, the symptoms of cholera begin with diarrhea and vomiting within 12–24 h. In healthy human volunteers, it was found that it takes approximately 106 bacterial cells to cause disease. However, a hypochlorohydric person can get cholera with a much smaller dose (Bennish, 1994). Severe dehydration due to diarrhea and vomiting causes sunken eyes and wrinkling of the abdominal skin and these symptoms are commonly used for visual diagnosis of the disease. Clinical diagnosis, including laboratory confirmation of V. cholerae infection, is not required for treatment since replacement of body fluid and electrolytes through intervenous injection or oral administration should be done quickly when treating cholera patients. Antibiotics, such as tetracycline, furazolidane, and doxycycline, can reduce the severity of the disease, reduce the volume of watery stool, and, thereby, the duration of illness (Bennish, 1994). Although modern molecular methods now enable microbiologists to gain new insight into the potential pathogenicity of environmental and clinical strains and to develop new strategies, especially in vaccine therapy, management of the disease continues to be based on urgent and appropriate replacement of fluids lost through diarrhea. Other than the oral or, in most severe cases, intravenous rehydration, recent studies emphasize the need for effective local health services to prevent infection. In fact, outside the context of endemic or seasonal epidemic cholera, the risk of acquiring the disease is low and can be reduced significantly by making good sanitation and safe potable water possible. However, because epidemic strains are indigenous to the aquatic environment, it is not possible to totally eliminate the cholera bacteria. However, as evidenced by the recent cholera outbreaks in Mozambique, South Africa, and Madagascar, resistance to tetracycline in V. cholerae has been increasing in the last decade (Dalsgaard et al., 1999; Duncan et al., 1994). Studies performed in central and east Africa (Materu et al., 1997) and south Asia (Steinberg et al., 2001) indicate that there is also an increase in multidrug resistance, but a lower prevalence of antimicrobial resistance in O139 serotypes compared with O1 serotypes (Ramamurthy et al., 2000). Resistance of V. cholerae O1 to fluoroquinolones has also been reported. Antimicrobial resistance genes in V. cholerae have been shown to be plasmid mediated and these may be transferred from resistant nonpathogenic gut coliforms in vivo to, and among, environmental strains (Faruque et al., 1998). Integron-borne gene cassettes within transferable plasmids, coding for multiple resistance genes, have been described in V. cholerae O1 strains isolated in Guinea-Bissau (Dalsgaard et al., 2000) and Vietnam (Dalsgaard et al., 1999). Thus, appropriate use of antimicrobials, understanding the molecular basis of antimicrobial resistance, and antimicrobial susceptibility testing of all V. cholerae clinical and environmental isolates must play a pivotal role in cholera surveillance and control programs (Shears, 2001; West, 1989).

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9.3. VIBRIOS AND THE AQUATIC ENVIRONMENT 9.3.1. Influence of Water Temperature, Salinity, and Nutrient Concentration Members of the genus Vibrio are natural inhabitants of the aquatic environment where they play important roles in nutrient cycling by remineralization of organic nutrients (Bassler et al., 1991; Chakraborty et al., 1997; Colwell, 1981). Vibrio concentrations in water are related to temperature, salinity, and presence of planktonic organisms. Most Vibrio spp. are characterized by increased growth at elevated temperatures, which is evident in the higher rates of isolation in the environment during warm months (Lipp et al., 2002) and in the presence of planktonic copepods (Huq et al., 1990). Until the late 1970s and early 1980s, V. cholerae was believed to be highly host-adapted and incapable of surviving longer than a few hours or days outside the human intestine. Since then, studies have revealed the existence of V. cholerae as free-living bacteria or in association with phytoplankton, zooplankton, crustaceans, and mollusks in coastal and estuarine environments (Chowdhury et al., 1997; Colwell & Huq, 1994; Janda et al., 1988; Montillo et al., 1995). These observations, and the capacity of the organism to adaptively respond to changes in salinity, temperature, and availability of nutrients (Huq et al., 1980), have shown that V. cholerae can successfully occupy a variety of aquatic habitats. Detection of V. cholerae, both O1 and non-O1 strains, has been reported in rivers and coastal waters in almost every part of the world (Huq et al., 2001b). However, non-O1 and non-O139 strains are more frequently isolated from rivers and estuarine areas than the O1 and O139 strains, and interestingly most environmental O1 strains that are isolated and cultured in the laboratory are nontoxigenic (Colwell & Spira, 1992), despite the fact that toxigenic strains can be identified in environmental samples using gene probes (Gil et al., 2004). It was hypothesized that nontoxigenic environmental strains can be converted by phage transduction with cholera toxin (CT)-encoding phage CTx yielding new detectable toxigenic strains (Faruque et al., 1998). Although its optimal salinity for growth is between 5 and 25‰ , V. cholerae is one of the few Vibrio spp. that can tolerate salinity of 0‰ (Singleton et al., 1982) and, in the presence of dissolved organic matter, it can also grow at salinities close to 45‰ (Singleton et al., 1982). Studies conducted by Miller et al. (1984) suggested that strains of V. cholerae vary greatly in their survival in the culturable state under low-salinity conditions (0.05‰ ). Seasonal disappearance of toxigenic strains of V. cholerae O1 both from human populations and from the environment in cholera endemic regions has been attributed to environmental factors related to changes in cellular physiology (Carroll et al., 2001; Fasano, 2000; Montilla et al., 1996; Singleton et al., 1982). The loss of culturability in this pathogen, leading to a dormant state, could be due, under most circumstances, to suboptimal growth temperatures and nutrient deprivation in aquatic systems (Adhikari, 1975; Singleton et al., 1982). As described below, plankton serves as a reservoir for this pathogen, providing the opportunity for it to exist in a dormant state (Singleton et al., 1982). Survival is not related to serogroup, source, or geographical origin of the strain (Miller et al., 1984). However, Montilla et al. (1996) have shown that the physiological state of the cells, and to a lesser degree, temperature and salinity are important factors in the conversion of V. cholerae from the non-O1 to the O1 serogroup. V. cholerae also grows under high-pH conditions, a parameter that has been used in enrichment for V. cholerae from environmental samples. V. cholerae remains stable in full sunlight, compared to enteric bacteria, such as E. coli, which may also impart selective advantage to vibrios in tropical latitudes (Mezrioui et al., 1995).

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V. parahaemolyticus grows well in media supplemented with 20–30‰ sodium chloride and preferentially inhabits brackish aquatic environments. Despite this, under specific nutrient conditions, sodium ion requirement is not mandatory and V. parahaemolyticus can survive well in areas where the salt concentration is lower than physiological concentrations (Sarkar et al., 1985). A detailed study of the seasonal distribution of V. parahaemolyticus showed that increases in isolation rate paralleled temperature rise (Sarkar et al., 1985). V. vulnificus has properties similar to those of V. parahaemolyticus, namely it is a normal habitant of marine and brackish water environments and results of many studies show that it can be isolated from water where the temperature is higher than 10 ◦ C (Oliver et al., 1983; Kelly, 1982; Cook, 1994; Chakraborty et al., 1997). V. mimicus is nonhalophilic and inhabits fresh or brackish environments; it is not found in aquatic environments with an average salinity ≥1%. In subtropical areas, it is isolated from the environment throughout the year, whereas in temperate waters it cannot be found in the winter (Chowdhury et al., 1989; Chakraborty et al., 1997). Bacteria are highly regulated to accommodate changing environments (Singleton et al., 1982; Weiner, 1999) and, under conditions of limited exogenous nutrient, can undergo a number of morphological and physiological changes (Roszak & Colwell, 1987; Tabor et al., 1981). Baker et al. (1983) reported that exposure of V. cholerae to nutrient deprivation caused the cells to become coccoid and lose over 90% of their original volume in 30 days, increase in cell number, lose small granules and inclusion bodies, lose the distinct three-layered integrity of the outer membrane, peptidoglycan, and the inner membrane, retaining only remnants of those structures, compress the nuclear region into the center of the cell surrounded by a denser cytoplasm, and form extended or convoluted structures from the cell wall which are pulled away from the cell membrane. It was suggested that these cell responses reflect the existence of strategies to enhance survival under conditions of exogenous nutrient deprivation (MacDonell & Hood, 1982; Stevenson, 1978). In the natural environment, bacteria with the greatest capacity to survive starvation transform into small spherical cells that develop various stress-resistance functions and have lower metabolic rates than nonstressed bacteria (Gras-Rouzet et al., 1996). Change from rod to ovoid morphology occurs in V. cholerae when it becomes nonculturable in response to nutrient deprivation, elevated salinity, and/or reduced temperature (Chaiyanan et al., 2001). Oliver et al. (1991) observed that when V. vulnificus was incubated in artificial seawater at 5 ◦ C, those cells decreased in size and changed from rods to cocci without increase in cell number. Along with reduction in size, the cells became undetectable by culture. Change in size and shape of bacteria subjected to starvation and suboptimal environmental conditions suggests smaller size is the natural state of bacteria in the environment (Byrd, 2000). In addition to a lower surface-to-volume ratio, smaller size can enhance survival by providing protection against predation and functioning as a response mechanism upon interaction with surfaces and at interfaces.

9.3.2. The Viable But Nonculturable State Vibrios and other non-spore-forming Gram-negative and Gram-positive bacteria exposed to environmental conditions (e.g. temperature, nutrient concentration) unfavorable for active growth and cell division, possess the ability to switch to a dormant stage, termed viable but nonculturable (VBNC). This phenomenon represents a state of dormancy and allows survival and persistence of bacterial cells in the natural environment (Lipp et al., 2002; Colwell & Huq, 1994; Huq et al., 2000; Munro & Colwell, 1996).

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In 1982, Xu et al. (1982) first demonstrated the VBNC phenomenon in E. coli and V. cholerae using the direct viable count and fluorescent-antibody staining methods. Cells in this state cannot be cultured on bacteriological media, yet maintain metabolic activity and the capability for re-growth upon exposure to appropriate stimuli, such as transition from the aquatic system, after ingestion, into the human intestinal tract (Colwell & Huq, 1994; Colwell et al., 1985, 1996; Gauthier, 2000; Oliver et al., 1991). Although molecular and immunological methods confirm viability of VBNC cells (Knight, 2000; Lle`o et al., 2000), the VBNC phenomenon continues to be debated in the scientific community, as the definition of living, as opposed to dead, bacteria is complex and difficult to articulate (Lipp et al., 2002). Some investigators simply state that bacterial cells not growing on culture plates are dead; yet Gonzalez et al. (1992) concluded that bacterial cells should be considered dead only when they lose both culturability and cellular integrity. Early studies by Valentine and Bradfield (1954), Postgate (1969), Jannasch (1969), Stevenson (1978), Kurath and Morita (1983), and others provide a framework for the concept of a resting stage for nonsporulating bacteria. Xu et al. (1982) concluded that cells unable to grow on conventional culture media but responsive to nalidixic acid treatment i.e., remaining metabolically active, are viable even when not culturable in the laboratory. Data from both laboratory experiments and field studies suggest that vibrios activate the VBNC state in response to either suboptimal temperature or nutrient deprivation or a combination of both, depending on species (Gauthier, 2000). The VBNC phenomenon is extensively described in different chapters of this book; however, a few reports on the potential pathogenicity of VBNC vibrios are briefly summarized as follows. The first data on this potential were obtained in V. cholerae and E. coli where accumulation of fluid in the ileal loop of rabbits after inoculation with VBNC cells was demonstrated (Colwell et al., 1985). Subsequently, a human volunteer study was carried out using VBNC cells of an attenuated strain of V. cholerae O1, and showed the development of clinical symptoms of cholera in one of two volunteers. However, the stools of both volunteers yielded culturable vibrios (Colwell et al., 1996), indicating that infection with VBNC bacteria can yield different results with different hosts. Furthermore, field studies conducted in Bangladesh, in which stool samples were collected from patients who were symptomatically suffering from cholera, showed that 8% of culture-negative stools turned out to be positive by culture-independent tests (Colwell, 2000). VBNC cells obtained by incubating V. cholerae O1 and O139 in laboratory microcosms prepared with 1% Instant Ocean at 4 ◦ C retained their antigenic determinants and their cholera toxin and toxin-associated genes, ctxA, toxR, tcpA, and zot (Chaiyanan et al., 2001). In another study, Pruzzo et al. (2003) showed that nonculturable V. cholerae O1 do not lose adhesive properties toward cultured intestinal cells after long term exposure to ASW. An initial study conducted by Linder and Oliver (1989) on V. vulnificus infectivity for mice failed to demonstrate virulence (Linder & Oliver, 1989). However, in a subsequent study, culturable cells of V. vulnificus, with identification confirmed using PCR, were recovered from blood and the peritoneal cavities of mice that had died from injections of VBNC cells which had been VBNC for at least 3 days (Oliver & Bockian, 1995). These data suggested that cells of V. vulnificus remain virulent, at least for some time, when in the VBNC state, and are capable of causing fatal infections following in vivo resuscitation. However, virulence decreases significantly as cells enter the VBNC state, which may account, at least to some extent, for the decrease in infections caused by this bacterium during winter months (Oliver & Bockian, 1995). The potential pathogenicity of the nonculturable cells was confirmed by semi-nested reverse transcription-PCR (RT-PCR) applied to VBNC populations of V. vulnificus

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prepared by exposure to 4 ◦ C in ASW (Fischer-Le Saux et al., 2002). By this assay, transcripts of the cytotoxin-hemolysin virulence gene vvhA were detected in nonculturable cells over a 4.5-month period, with a progressive decline in signal over time. Pace et al. (1997) showed that bile triggers release from dormancy in V. parahaemolyticus and enhances expression of virulence factors. In another study, VBNC forms of V. alginolyticus and V. parahaemolyticus environmental strains and V. parahaemolyticus ATCC 43996 showing virulence characteristics (hemolysin production, adhesive and/or cytotoxic ability, in vivo enteropathogenicity) were obtained by culturing bacteria in a microcosm consisting of ASW, at 5 ◦ C (Baffone et al., 2003). The VBNC forms were resuscitated in a murine model. The virulence characteristics that seemed to disappear after resuscitation in the mouse were subsequently reactivated by two consecutive passages of the strains in the rat ileal loop model. It is concluded from these studies that, in the VBNC state, pathogens maintain virulence genes and virulence potential; moreover, in vivo passage can trigger VBNC cells to revert to the culturable, virulent state. The fact that VBNC cells maintain adhesive properties, although at a lower degree than culturable bacteria, suggests that when dormant bacteria present in the environment are ingested via contaminated food or water and reach the intestinal tract, they may persist on the intestinal epithelium. If reactivation occurs, adhering vibrios may fully express their pathogenic potential. Therefore, these observations emphasize the need for reliable molecular (e.g. PCR, RT-PCR) and immunological methods for detecting VBNC bacteria in both environmental and clinical samples in order to reduce the risk of false negative results and the potential for transmission to humans.

9.3.3. Interactions with Living and Nonliving Substrates Present in Seawater 9.3.3.1. Biofilm Formation A common feature among vibrios is the presence of multiple lifestyles: a planktonic, free swimming state and a sessile existence on a surface, forming structurated communities called “biofilms.” A biofilm may be developed in a multistep process, including random collision or chemotaxis, reversible and irreversible attachment, microcolony formation, and ordered threedimensional structure formation. This structure is composed of pillars of bacteria surrounded by water channels that allow nutrients to reach biofilm-associated bacteria and allow toxic metabolites to diffuse out of the biofilm (Costerton et al., 1995). Multiple studies have been conducted to understand the mechanisms by which vibrios and, in particular V. cholerae, adhere to different surfaces, inside and outside the human host. In general, the adhesive process appears to be multifactorial, with contributions from a variety of different cell surfaces and secreted components. In the aquatic environment, in consideration of their chitinolytic activity, attention has been given to chitin containing surfaces. The early studies performed using V. furnissii showed that populations of cells exhibit an adhesion–deadhesion behavior that allows free-swimming bacteria to relocate when a chitin source is depleted and to look for another source (Yu et al., 1991). Bassler et al. (1991) subsequently demonstrated that vibrio cells are also positively chemotactic to the products of chitin hydrolysis. Surface proteins mediating attachment to chitin particles have been identified in Vibrio harveyi (Montgomery & Kirchman, 1993, 1994; Yu et al., 1991), V. alginolyticus (Pruzzo et al., 1996), and V. cholerae O1 classical strains (Tarsi & Pruzzo, 1999). During exposure in ASW for relatively long periods of time at 5 ◦ C and 18 ◦ C, V. cholerae O1 retains

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the ability to express chitin-specific ligands and to interact with chitin particles and copepods, though less efficiently than when cells are incubated at 35◦ C (Pruzzo et al., 2003). MSHA pilus, which is present in V. cholerae strains of either El Tor biotype or O139 serogroup (Jonson et al., 1994; Tacket et al., 1998), has been shown to be involved in attachment to nonnutritive surfaces (borosilicate and cellulose) (Watnick et al., 1999) and in colonization of exoskeletons of the planktonic crustacean Daphnia pulex (Chiavelli et al., 2001). According to the model proposed for V. cholerae biofilm formation (Bomchil et al., 2003; Watnick et al., 2001; Watnick & Kolter, 1999), bacteria first swim toward the surface using their polar flagella and establish cell-to-surface interactions. These interactions are greatly enhanced by the presence of MSHA pilus. Once contact with the surface has been made, bacteria move along the surface by means of their flagella, recruit additional planktonic cells, and divide, resulting in the formation of microcolonies. Attached cells synthesize an exopolysaccharide, produced by the vps genes, which is thought to be a major component of the extracellular matrix that stabilizes the mature biofilms forming the three-dimensional architecture. However, Kierek and Watnick (2003) showed that V. cholerae forms vps-dependent and vps-independent biofilms; environmental activators of vps-dependent biofilm development are present in freshwater, while environmental activators of vps-independent biofilm development are present in seawater. Recently, a novel gene named mbaA (for maintenance of biofilm architecture) has been reported to play a role in the formation and maintenance of the highly organized threedimensional architecture of V. cholerae El Tor biofilms (Bomchil et al., 2003). The predicted mbaA product is a member of a family of regulatory proteins, suggesting that MbaA regulates the synthesis of some component of the biofilm matrix. Expression of the exopolysaccharide causes a rugose colony phenotype, and provides enhanced chlorine resistance and also phage resistance (Reidl & Klose, 2002). Interestingly, exopolysaccharide expression inhibits intestinal colonization in the infant mouse cholera model (Watnick et al., 2001); on the other hand, MSHA pilus that is required to form a normal biofilm is not directly involved in adhesion to intestinal epithelial cells (Attridge et al., 1996; Kaper et al., 1995; Tacket et al., 1998), suggesting that the expression of factors that enhance environmental persistence is not related to virulence. V. parahaemolyticus possesses multiple cell types appropriate for life under different circumstances (Belas et al., 1986; McCarter, 2000, 2001; McCarter & Silverman, 1990; McCarter et al., 1988). The swimmer cell, with a single polar flagellum, is adapted to life in liquid environments. The polar flagellum, powered by the sodium motive force, is constitutively expressed and propels the bacterium at various speeds, allowing it to find and closely approach a surface. The swarmer cell, propelled by many proton-powered lateral flagella, can move through highly viscous environments, colonize surfaces, and form multicellular communities. Lateral flagella are encoded by one of the two distinct flagellar gene sets in V. parahaemolyticus; each set contains more than 35 structural and regulatory genes. Signals that induce differentiation to the surface-adapted cell type are both physical and chemical in nature (McCarter, 2001; Stewart & McCarter, 2003). Genetic rearrangements create additional phenotypic versatility for the organism and is manifested as variable opaque and translucent colony morphotypes (McCarter, 2000). Switching between the two morphotypes may occur randomly or it may be responsive to specific environmental conditions. The discovery that a LuxR homolog controls the opaque cell type implicates inter-cellular signaling as an additional survival strategy. The alternating identities of V. parahaemolyticus may play important roles in attachment and detachment, as bacterial population adapt to growth on surfaces, to form structurated communities and develop biofilms.

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9.3.3.2. Vibrios and Planktonic Organisms In the aquatic ecosystem chitin is the most abundant polysaccharide and the principal component of zooplankton exoskeleton. Chitinous organisms, e.g. copepods, amphipods, and other crustaceans, are prevalent among zooplankton populations. Approximately 1011 metric tons of chitin/year is produced in aquatic environments and >109 metric tons is produced by copepods alone (Keyhani & Roseman, 1999). Marine copepod exoskeleton has been shown to support large population of vibrios, including pathogenic species as V. cholerae, V. parahaemolyticus, V. vulnificus, and V. alginolyticus (Carli et al., 1993; Colwell, 1981; Colwell & Huq, 1999; Hansen & Bech, 1996; Huq et al., 1983; Kaneko & Colwell, 1975; Tamplin et al., 1990). Without bacterial activity that returns insoluble polysaccharide to the ecosystem in a biologically useful form, seawater would become depleted of carbon and nitrogen in relatively short time (Yu et al., 1991). The commensal relationship between vibrios and copepods has important consequences (Colwell, 1981, 1996): (i) bacteria bound to exoskeleton are able to metabolize the chitin more efficiently than free-living forms, thus increasing the rate of chitin mineralization in the natural environment; (ii) attachment to copepods and, in particular, to the eggs that are dispersed in the water provides a mechanism for extended geographic distribution of bacteria, including pathogens; (iii) plankton organisms are the most plausible reservoir of altered forms of pathogens, e.g. V. cholerae in the VBNC stage (Barcina et al., 1997; Colwell, 1981; McDougald et al., 1998), from which fully virulent strains can arise; and (iv) epibiotic bacteria are protected against environmental stress and thrive on excreted products of digestion, including organic material and NH+ 4 (Cooksey & Wigglesworth-Cooksey, 1995). V. cholerae has also been found to be episodically associated with a wide range of marine and freshwater living organisms (Hood & Winter, 1997) including cyanobacteria (Anabaena variabilis) (Islam et al., 1989), diatoms (Skeletonema costatum) (Martin & Bianchi, 1980), filamentous green algae (Rizoclonium fontanum) (Islam et al., 1989), water hyacinths (Eichornia crassipes) (Spira et al., 1981), arthropods (Gerris spinolae) (Shukla et al., 1995) and blue crabs (Callinectes sapidus) (Shukla et al., 1995). Recently, environmental non-O1 and non-O139 V. cholerae strains were isolated from chironomid egg masses (Chironomus sp., Diptera) collected from waste stabilization ponds (Broza & Halpern, 2001); it was suggested that these egg masses may serve as a favorable rich nutrient niche for V. cholerae strains under certain conditions, e.g., polluted water. Recent evidence indicates that there may be a specific symbiosis between species of copepods and vibrios (Lipp et al., 2002); moreover, scanning electron microscopy studies have shown that bacterial attachment to exoskeleton is not a random process. Egg sacs and the oral region of copepods are sites of significantly enhanced attachment of V. cholerae (Dumontet et al., 1996; Huq et al., 1983); in contrast, V. parahaemolyticus was found to bind uniformly to the entire surface of the copepod exoskeleton (Huq et al., 1983). Cells of V. cholerae also colonize the gut of copepods, and Singleton et al. (1982) proposed that there might be an environmental role for CT production in this type of association. As shown in human intestine, cholera toxin might affect efflux of Na+ , other electrolytes, and water of epithelial cells of crustacea thereby providing an effective mechanism to regulate the concentration of Na+ when it is not optimal for the host (Lipp et al., 2002). Interactions between vibrios and copepods are affected by environmental parameters (Kaneko & Colwell, 1975). For instance, increased salinity decreases the numbers and rate

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of attachment of V. parahaemolyticus. In a sample of estuarine waters with a salinity of 2‰ , all V. parahaemolyticus cells were found to bind purified chitin; however, the percentage of adsorbed bacteria decreased linearly to 70% when the salinity was increased to 16‰ . This trend was also observed with V. parahaemolyticus binding to chitin in NaCl solutions, and binding to copepods in natural seawater/NaCl solutions (Koneko & Colwell, 1975). V. alginolyticus adsorption reaches 80% in the presence of 30‰ NaCl and 4‰ MgSO 4 , MgCl2 , and CaCl2 , at 20 ◦ C with pH ranging from 7 to 8 (Pruzzo et al., 1996). A salinity of 15‰ and temperature ranging from 25 to 30 ◦ C were found to positively influence attachment of V. cholerae to copepods (Huq et al., 1984). There may be a synergistic effect between phytoplankton and zooplankton in chitin colonization by V. cholerae: an alkaline pH of 8.5, often associated with algal blooms, was found to positively influence the attachment of V. cholerae to copepods (Huq, 1984). Several data suggest that the surface and gut of zooplankton comprise an ecosystem that may influence the onset of the nonculturable state or initiate growth of these bacteria (Huq et al., 1983). V. cholerae and V. parahaemolyticus adhering to copepods are able to tolerate temperature shifts better than nonassociated bacteria (Huq et al., 1984). V. cholerae adhering to living copepods remained culturable 10 days longer than V. cholerae associated with dead copepods, indicating that living copepods confer some form of protection (Huq et al., 1984). Living copepods in microcosms were also able to protect bacteria from the stress of lower salinity (85% 80% 86–94% 94% 60% 35% 69% 93% 60% 43%

Data compiled from cases described by Oliver (1989), Blake et al. (1979), and Tacket et al. (1984), from data on 274 V. vulnificus cases reported by the US Food and Drug Administration for the period 1989–2000, and from reviews by both Oliver and Kaper (2001) and Strom and Paranjpye (2000).

as 7 h post-ingestion of oysters, to 10 days (Oliver, unpublished). Table 10.3, a compilation of over 300 cases described by Oliver (1989), Blake et al. (1979), and Tacket et al. (1984), data for the 274 V. vulnificus cases reported by the FDA for the period 1989–2000, and reviews by both Oliver and Kaper (2001) and Strom and Paranjpye (2000), shows the symptoms typical of V. vulnificus primary septicemia. The most commonly reported symptoms are fever (94%), nausea (60%), and hypotension (systolic blood pressure 10 (and often >108 ), whereas the opaque morphotype yields LD50 values of 105 CFU/g, whereas others had undetectable levels (Birkenhauer & Oliver, 2003). Such variations in load are typical (Cook, 1994; Cook & Ruple, 1989, 1992; DePaola et al., 1994; Tamplin, 1994). Studies on the presence of V. vulnificus in shellfish, primarily oysters, have also indicated a strong correlation with temperature; in general, investigators are unable to isolate V. vulnificus from oysters taken from cold waters. This finding also correlates with epidemiological data indicating low numbers of oyster-associated human cases during winter months (November through March in the United States).

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V. vulnificus (cfu per g oyster)

1.E+06

1.E+05

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1.E+03

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1.E+02 1

2

3

4

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6

7

8

9

10

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oyster Figure 10.4. Levels of V. vulnificus among individual oysters in a batch. Asterisk indicates levels below the limit of detection (202 µm class) were studied, levels of 101 –104 per cell were found. In a study of goby taken from Tokyo Bay, Aono et al. (1997) reported 0.3% of the 1056 vibrios they isolated from this fish to be V. vulnificus, based on PCR analysis. DePaola et al. (1994) conducted what is probably the most extensive study on the presence of V. vulnificus in a wide variety of fish. They studied the occurrence of this species in 23 different fish species taken from coastal and offshore waters of the Gulf of Mexico. Of the nine off-shore species, only two yielded V. vulnificus, consistent with the generally reported inability to isolate this bacterium in open-ocean waters. However, they did isolate V. vulnificus, often in very high numbers, from all coastal fish studied. They were also able to isolate V. vulnificus from fish taken during the winter months, when levels in surrounding water and sediments were undetectable or extremely low. They thus theorized that fish might represent a

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reservoir during cold-water months, when V. vulnificus was frequently isolated at high densities from such bottom-feeding fish such as sheepshead. Levels as high as 106 /g were found in the intestinal contents of some bottom-feeding fish, especially those that consume mollusks and crustaceans. The finding that sediment-associated animals may harbor V. vulnificus during cold-water months is in support of an earlier study by Vanoy et al. (1992). They were able to isolate V. vulnificus from sediments during cold-water months, but were unable to isolate this species from water, suspended particulate matter, or oysters. They thus suggested that V. vulnificus may also “over-winter” in sediments. The occurrence of V. vulnificus in sediments has been reported by relatively few investigators. Vanoy et al. (1992) reported V. vulnificus to range from 105 /g of sediments taken from Galveston Bay in Texas. Similarly, a study by Williams and LaRock (1985) reported concentrations from 104 /g, with 50% of the samples yielding 100 or fewer V. vulnificus cells per gram. Aono et al. (1997) took over 2600 sediment samples from Tokyo Harbor and found that while vibrios compised 20% of the resident microflora, only 0.2% could be identified as V. vulnificus. Hø i et al. (1998) were only able to isolate V. vulnificus from ∼57% of the sediment samples they examined, and Pfeffer et al. (2003) could only confirm a single isolate of this species in sediments over a 6-month (cold-water) period. Thus, not only the levels of this species in estuarine sediments appears to vary considerably, but also the water samples appear as likely to equal or exceed the number of cells found in sediments.

10.7. FACTORS AFFECTING DISTRIBUTION IN THE ENVIRONMENT 10.7.1. Salinity The two most significant factors affecting the ecology of V. vulnificus in seawater are water temperature and salinity. As noted above, the bacterium is an obligate halophile, and is found only in waters with salinities of at least 5‰ . The bacterium does not appear to survive well in elevated salinities, has not been isolated from open-ocean waters, and most studies isolate this species from water with salinity values between 8‰ and 23‰ (e.g. DePaola et al., 1994; Høi et al., 1998; Tamplin et al., 1982). In a study of 21 sites in the Gulf of Mexico (Kelly, 1982), V. vulnificus was recovered from all waters with relatively low (7–16‰ ) salinities. Oliver et al. (1983) also reported a negative correlation with salinity. In a study of oysters taken from Gulf and Atlantic Coast waters (Motes et al., 1998), salinity was found to be a controlling factor for the levels of V. vulnificus in these shellfish, with highest numbers (>103 /g) found in waters of intermediate salinities (5–25‰ ).

10.7.2. Temperature Virtually all studies have found water temperature to be a key factor in the ability to isolate V. vulnificus (Kelly, 1982; Høi et al., 1998; Jones & Summer-Brason, 1998; Tamplin et al., 1982; Tilton & Ryan, 1987; Vanoy et al., 1992; Veenstra et al., 1994), and most investigators report isolation of this species only when water temperatures are between 15◦ C and 32◦ C. Investigators sampling regions that typically have low water temperatures are often unable to isolate V. vulnificus (Kaysner et al., 1990). Figure 10.5 shows the strong correlation between water temperature and V. vulnificus levels we found during a multi-year study of six estuarine areas in the eastern United States (Pfeffer et al., 2003). As is evident, the levels of this pathogen

268 102

30 25

101 20 100

15 10

10−1

Temperature (°C)

V. vulnificus (cfu per g oyster)

J. D. Oliver

5 10−2

0 0

5

10

15

20

Temperature V. vulnificus

25

Month (7/00 4/02) Figure 10.5. Correlation between water temperature and levels of V. vulnificus in seawater. Taken from Pfeffer et al. (2003). Reprinted with permission from the American Society for Microbiology, Washington, DC.

decreased to undetectable levels when water temperatures decreased below 15◦ C. Indeed, following a multiple regression analysis of the data, we found water temperature to account for most (47%) of the variability in concentration of V. vulnificus in these estuarine waters. This is similar to the level (65%) reported by Motes et al. (1998) for the isolation frequency of V. vulnificus from oysters. This correlation between the ability to isolate V. vulnificus and water temperature is also paralleled by the incidence of human infection (85% of the cases described in Fig. 10.2

V. vulnificus (cfu per g oyster)

1.E+05

1.E+04

1.E+03

1.E+02

1.E+01

* February

* March

April

June

July

August

September

October

month Figure 10.6. Seasonal variation of V. vulnificus in oysters harvested from the Gulf Coast of the United States.V. vulnificus levels are indicated as CFU/g of oyster tissue. Asterisk indicates levels below the limit of detection (200 µm) than intermediate size plankton (ranging from 64 to 200 µm). Proof of the authentic binding of whole enterococcal Table 13.2. Comparison of numbers of positive samples of enterococci detection collected along side the Italian cost during years 2001–2002. Sampling site Ancona (Adriatic Sea) Pescara (Adriatic Sea) Pesaro (Adriatic Sea) Genova (Ligurian Sea) Cagliari (Tyrrhenian Sea) Sassari (Tyrrhenian Sea) Messina (Tyrrhenian Sea)

Total samples no.

Positive by culture (%)

Positive by PCR (%)

51 51 45 39 36 33 36

5 (9.80) 5 (9.80) 3 (6.66) 3 (7.69) 6 (16.60) 0 (0.00) 15 (41.66)

28 (54.90) 36 (70.58) 24 (53.33) 19 (48.71) 17 (47.22) 14 (42.42) 26 (72.22)

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Table 13.3. Seasonal distribution of enterococci detected by PCR from water and plankton fractions collected in the Ancona sampling station. Number of positive samples/total (%) in Seasona

Water

Plankton (64–200 µm)

Plankton (>200 µm)

Winter Spring Summer Fall

0/17 (0.00) 0/17 (0.00) 1/17 (5.88) 0/17 (0.00)

1/17 (5.88) 3/17 (17.64) 4/17 (23.53) 2/17 (11.76)

2/17 (11.76) 4/17 (23.53) 7/17 (41.18) 4/17 (23.53)

a

Winter: January, February, March; Spring: April, May, June; Summer: July, August, September; fall: October, November, December.

Table 13.4. Occurrence of enterococci in seawater and plankton fractions collected in the Ancona sampling station. Number of positive samples/total (%) detected by Sample Water Plankton (64–200 µm) Plankton >200 µm

Culture

PCR

0/51 (0.00) 2/51 (3.92) 3/51 (5.88)

1/51 (1.96) 10/51 (19.61) 17/51 (33.33)

cells and not simply of extracellular DNA as a result of cell lysis, is provided in Figure 13.3, which shows that anti-D antibodies coupled to microbeads are bound to copepod surface present in a sample that proved negative for enterococcal detection with the culture method and positive with PCR, while anti-D antibodies do not coat the surface of copepods from a sample found to be negative for enterococci by both the culture and molecular methods. This result is reminiscent of what has previously been described in Vibrio spp. (see also Chapter 9 of this book). Thus, enterococci, like many other pathogens, when dispersed in a non-friendly environment, bind to other

A

B

Figure 13.3. Microscopic observation of marine zooplankton treated with anti-D antibody bound to latex microbeads. (A) Plankton sample negative for enterococci detection by culture but positive by PCR, and (B) a plankton sample negative for enteroccocci detection by both culture and PCR. Black arrows in (A) indicate the presence of latex microbeads as a black striation on the carapace. Empty arrows in (B) indicate the carapace without adherent microbeads.

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organisms instead of living in a planktonic phase. Aggregation of bacteria may be considered a more successful strategy in terms of bacterial survival. At present, the mechanism governing binding to marine plankton is not known, but either some form of specific binding can be postulated (a model drawn from Vibrio spp.) or we may simply consider plankton (and other marine micro- and macro-organisms) as a sort of sponge that traps enterococci inside itself. What is important to stress is the role played by marine organisms in concentrating pathogenic bacteria and, possibly, move them with currents and tides.

13.5. SIGNIFICANCE OF THE PRESENCE OF GRAM-POSITIVE BACTERIA IN SEAWATER: EPIDEMIOLOGY AND RISK FOR HUMAN HEALTH A substantial amount of epidemiological data has been accumulated on the relationship between bathing in microbiologically polluted water and illness. Many of these studies were conducted on seawaters collected near beaches of England, Spain, France, Israel, Hong Kong, and the United States (Cabelli et al., 1982; Brown et al., 1987; Kay et al., 1994), and the findings have recently been summarized by Pruss (1998) and Kay and Dufour (2000). All these studies have shown a higher rate of minor diseases among bathers in comparison with non-bather control groups. The diseases among swimmers detected in these studies were associated with the gut, ear, nose, throat, and skin. For the marine beaches, enterococci (measured at the chest depth) showed the closest relationship with gastroenteritis, though a good correlation was also observed in freshwaters for both enterococci and E. coli. Fleisher et al. (1996) identified a dose–response relationship between bathers exposed to seawater contaminated by domestic sewages and risk of non-enteric illness in the UK. The results indicate that exposure to fecal streptococci was predictive of acute febrile respiratory illness (AFRI) at levels of 60 fecal streptococci/100 ml or above. Prospective epidemiological studies of swimmers and non-swimmers were conducted on several US beaches, including both nonpolluted and polluted ones, with various measured concentrations of microorganisms. On the basis of these results, the USEPA, in 1986, established new microbiological criteria, indicating as maximum limit values a geometric mean of 126 E. coli per 100 ml and 33 and 35 enterococci per 100 ml of freshwater and seawater, respectively. This recommendation is based on the analysis of five water samples collected from a given swimming site over a 30-day period. The recommended standard correlates with an expected rate of 19 diarrheal diseases per 1000 swimmers. These limit values have not universally accepted. Canada, for an example, retained its older standard of 200 fecal coliforms per 100 ml. The European Community currently uses guidelines (Council Directive 76/160EEC), which state that 80% of the samples should not exceed 2000 total coliforms per 100 ml or 100 fecal coliforms and enterococci, both per 100 ml. The European guidelines are now being reassessed, but higher values for both fecal coliforms and enterococci are expected, namely, 200 CFU/100 ml for both enterococci and E. coli. The proposed standards correspond to a 5% risk of contracting gastroenteritis (Kay et al., 1994). Although several investigators have indicated the close relationship between enterococci and the risk of diseases, other microorganisms have been reported as alternatives. This is the case, among Gram-positive bacteria, of staphylococci, which have been indicated by some authors as being the best overall health risk indicator in several epidemiological studies in both the marine and freshwater environments (Seyfried et al., 1985; Cheung et al., 1991).

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As far as Listeria is concerned, though its detection is feasible in seawaters, no clear etiological link has ever been demonstrated as regards human infection. In addition, intriguing aspect of the microbiological significance of Gram-positive organisms in seawater, has to do with M. marinum and its role in human infection via fish-tank hobby and fish/shellfish injuries (Aubry et al., 2002). A retrospective survey demonstrates an unmistakable relationship between the presence of this microorganism and the rise in the number of pathologies occurring in several body districts.

13.6. CONCLUDING REMARKS AND FUTURE PROSPECTS Data from recently performed experiments unequivocally suggest that a number of pathogenic bacteria are able to persist in the extracorporal environments longer than was previously believed. The cause of this erroneous belief is essentially linked to the preconception that live bacteria are necessarily culturable in the laboratory on appropriate microbiological media. It is now clear that the medically important bacteria, including Gram-positive organisms, may persist in environments as starved, injured, VBNC cells; therefore, culture methods alone are no longer considered sufficient for the satisfactory protection of human health. The inadequacy of cultural methods has been clearly exposed by an experiment conducted by Colwell et al. (1996) in which human volunteers were fed with water containing VBNC V. cholerae cells. The volunteers manifested cholera symptoms and live V. cholerae cells were isolated from their stools. The obvious explanation of this result is that dormant VBNC cells were resuscitated in the gut of the volunteers. The crux of the matter then is to understand at what point a bacterium can be considered viable, or in other words, up to what point in its permanence in the environment is still capable of infecting human subjects and causing disease. The task therefore is to identify, for each microorganism that persists in a stress-inducing environment such as seawater, what can be called the point of no return, after which the bacterium can be considered dead, or at any rate incapable of resuscitating and causing infection in humans (see Fig. 13.4). To achieve this goal, the study of the physiology and the gene expression of VBNC cells is essential. Thus, specific targets suggestive of cell vitality need to be sought in order to devise protocols for the molecular detection of pathogenic and indicator bacteria in the environment. Messenger RNA, due to its short half-life, would at present seem to be a candidate for this role. In addition, particular attention needs to be devoted to the validation of the protocols regarding the standardization of extraction and purification of nucleic acid, due to the presence in environmental samples of inhibitors of both the retro-transcription and amplification reactions that could significantly reduce the sensitivity of the detection method. The establishment of refined detection methods will be facilitated by the recent availability of sophisticated methods appearing on the horizon of technology as applied to medical science and which are now beginning to make a practical impact, such as µSERS biochip, oligonucleotide microarray, and surface acoustic wave device biosensor (Bej, 2003). Another point warranting particular attention is the observation that enterococci mainly persist in a state adherent to plankton. Because marine organisms are able to concentrate bacteria and move them with currents and tides, this phenomenon should be investigated more thoroughly and its impact on human health assessed. As soon as the fate, state of activity and pathogenic potential of the bacteria discussed here have been properly assessed and appropriate molecular methods for their accurate detection and activity assessment have been developed, validated, and compared with classical methods;

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Figure 13.4. Schematic representation of the fate of human pathogens in the environment.

additional epidemiological studies will be needed to establish the relationship between the presence of stressed microbial forms in seawaters and the incidence of diseases in swimmers; all this with a view to establish reliable guidelines and to choose useful indicators in order to ensure effective risk management.

ACKNOWLEDGMENTS The authors wish to thank the members of research units participating in the Cofin2000 project on the development of new methods for the microbiological evaluation of freshwater and seawaters, and particularly W. Baffone (University of Urbino), L. Cellini (University of Chieti), M. T. Fera (University of Messina), L. Pane (University of Genoa), R. Pompei (University of Cagliari), C. Pruzzo (University of Ancona), and P. Rappelli (University of Sassari). Special thanks to the Ph.D. students D. Benedetti, B. Bonato, G. Burlacchini, and S. Pierobon whose work was essential for achieving the results described here. The technical assistance of M. C. Tafi and V. Marconi was also greatly appreciated. The studies described here were supported by grants from the Italian National Research Council (Targeted Project on Biotechnology), the Ministry of the University and Scientific Research (Cofin2000 and Cofin2003), and the University of Verona (Fondi ex60%).

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Kjelleberg (ed), Starvation in Bacteria. Plenum Press, New York, pp. 103–123. Paludan-Muller, C., Weichart, D., McDougald, D., & Kjelleberg, S. (1996). Analysis of starvation conditions that allow for prolonged culturability of Vibrio vulnificus at low temperature. Microbiology 142, 1675–1684. Pettibone, G.W., Sullivan, S.A., & Shiaris, M.P. (1987). Comparative survival of antibiotic-resistant and -sensitive fecal indicator bacteria in estuarine water. Appl Environ Microbiol 53, 1241–1245. Pommepuy, M., Butin, M., Derrien, A., Gourmelon, M., Colwell, R.R., & Cormier, M. (1996). Retention of enteropathogenicity by viable but nonculturable Escherichia coli exposed to seawater and sunlight. Appl Environ Microbiol 62, 4621–4626. Pruss, A. (1998). Review of epidemiological studies on health effects from exposure to recreational water. Int J Epidemiol 27, 1–9. Pruzzo, C., Tarsi, R., Lle`o, M.M., Signoretto, C., Zampini, M., Colwell, R.R, & Canepari, P. (2002). In vitro adhesion to human cells by viable but nonculturable Enterococcus faecalis. Current Microbiol 45, 105–110. Pruzzo, C., Tarsi, R., Lle`o, M.M., Signoretto, C., Zampini, M., Pane, L., Colwell, R.R., & Canepari, P. (2003). Persistence of adhesive properties in Vibrio cholerae after long-term exposure to sea water. Environ Microbiol 5, 850–858. Pyle, B.H., Broadaway, S.C., & McFeters, G.A. (1995). A rapid, direct method for enumerating respiring enterohemorrhagic Escherichia coli O157:H7 in water. Appl Environ Microbiol 61, 2614–2619. Resnick, I.G., and Levin, M.A. (1981). Assessment of bifidobacteria as indicators of human fecal pollution. Appl Environ Microbiol 42, 433–438. Rollenhagen, C., Antelmann, H., Kirstein, J., Delumeau, O., Hecker, M., & Yudkin, M.D. (2003). Binding of σA and σB to core RNA polymerase after environmental stress in Bacillus subtilis. J Bacteriol 185, 35–40. Rorvik, L.M., Caugant, D.A., & Yndestad, M. (1995). Contamination pattern of Listeria monocytogenes and other Listeria spp. in a salmon slaughterhouse and smoked salmon processing plant. Int J Food Microbiol 25, 19–27. Roszak, D.B., & Colwell, R.R. (1987). Survival strategies of bacteria in the natural environment. Microbiol Rev 51, 365–379. Schultz, J.E., & Matin, A. (1991). Molecular and functional characterization of a carbon starvation gene of Escherichia coli. J Mol Biol 218, 129–140. Scoglio, M.E., Di Pietro, A., Mauro, A., Piperno, I., Lagana, P., & Delia, S.A. (2000). Isolation of Listeria spp., Aeromonas spp., and Vibrio spp. from seafood products. Annali di Igiene 12, 297–305. Seyfried, P.L., Tobin, R.S., Brown, N.E., & Ness, P.F. (1985). A prospective study of swimming-related illness. II. Morbidity and the microbiological quality of water. Am J Public Health 75, 1071–1075. Sheridan, G.E., Masters, C.I., Shallcross, J.A., & MacKey, B.M. (1998). Detection of mRNA by reverse transcriptionPCR as an indicator of viability in Escherichia coli cells. Appl Environ Microbiol 64, 1313–1318. Signoretto, C., Boaretti, M., & Canepari, P. (1994). Cloning, sequencing and expression in Escherichia coli of the low-affinity penicillin binding protein of Enterococcus faecalis. FEMS Microbiol Lett 123, 99–106.

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Signoretto, C., Boaretti, M., & Canepari, P. (1998). Peptidoglycan synthesis by Enterococcus faecalis penicillin binding protein 5. Arch Microbiol 170, 185–190. Signoretto, C., Burlacchini, G., Lle`o, M.M., Pruzzo, C., Pane, L., Franzini, G., & Canepari, P. (2004). Adhesion of Enterococcus faecalis in the nonculturable state to plankton is the main mechanism responsible for persistence of this bacterium in both lake and seawater. Appl Environ Microbiol 70, 6892–6896. Signoretto, C., Lle`o, M.M., Tafi, M.C., & Canepari. P. (2000). Cell wall chemical composition of Enterococcus faecalis in the viable but nonculturable state. Appl Environ Microbiol 66, 1953–1959. Soontharanont, S., & Garland, C.D. (1995). The occurrence of Listeria in temperate aquatic habitats. Proceedings of XII International Symposium on problems of Listeriosis. Perth, Western Australia, Publication Promaco Conventions, pp. 145–146. Takayama, K., & Kjelleberg, S. (2000). The role of RNA stability during bacterial stress responses and starvation. Environ Microbiol 2, 355–365. Toma, F. (1998). Nontuberculous mycobacteria in the hospital environment. Bacteriol Virusol Parazitol Epidemiol 43, 229–235. Trainor, V.C., Udy, R.K, Bremer, P.J., & Cook, G.M. (1999). Survival of Streptococcus pyogenes under stress and starvation. FEMS Microbiol Lett 176, 421–428. Vives-Rego, J., Lebaron, P., & Nebe-von Caron, G. (2000). Current and future applications of flow cytometry in aquatic microbiology. FEMS Microbiol Rev 24, 429–448. Whitman, R.L., Shively, D.A., Pawlik, H., Nevers, M.B., & Byappanahalli, M.N. (2003). Occurrence of Escherichia coli and enterococci in Cladophora (Chlorophyta) in nearshore water and beach sand of Lake Michigan. Appl Environ Microbiol 69, 4714–4719. Wiggins, B.A., Andrews, R.W., Conway, R.A., Corr, C.L., Dobratz, E.J., Dougherty, D.P., Eppard, J.R., Knupp, S.R., Limjoco, M.C., Mettenburg, J.M., et al. (1999). Use of antibiotic resistance analysis to identify nonpoint sources of fecal pollution. Appl Environ Microbiol 65, 3483–3486. Wittwer, C.T., Herrmann, M.G., Moss, A.A., & Rasmussen, R.P. (1997). Continuous fluorescence monitoring of rapid cycle DNA amplification. Biotechniques 22, 134–138. Yoshpe-Purer, Y., & Golderman, S. (1987). Occurrence of Staphylococcus aureus and Pseudomonas aeruginosa in Israeli coastal water. Appl Environ Microbiol 53, 1138–1141.

14 Fecal Contamination in Coastal Areas: An Engineering Approach M. Pommepuy, D. Hervio-Heath, M. P. Caprais, M. Gourmelon, J. C. Le Saux, and F. Le Guyader 14.1. INTRODUCTION “The occurrence of pathogenic microorganisms in seawater or in shellfish could exist anytime sewage from human or animal origin would be discharged to the coast” (Metcalf, 1982). According to the diseases occurring in the human population or in animals, pathogens might be present in recreational waters or in shellfish. Thus, the presence of human enteric viruses (norovirus, astrovirus, rotavirus, hepatitis A virus (HAV)) and pathogenic bacteria (Salmonella, Listeria monocytogenes, Shiga-toxin-producing Escherichia coli (STEC), Vibrio cholerae, Vibrio parahaemolyticus, etc.) has been reported in coastal areas for a long time (Colwell, 1978; Metcalf, 1978; Melnick et al., 1979; Grimes, 1991; Bosch et al., 2001; Kong et al., 2002). These microorganisms have been implicated in gastrointestinal and respiratory illnesses and other infections (skin, eyes, etc.), (Griffin et al., 2003). Using risk-assessment models for viruses, maximum risks were estimated to be 1.3 infections per 100 swimmers (Colwell et al., 1996). To evaluate the risk due to the presence of these pathogens in the environment, certain criteria have to be determined. Among them, the infectious dose would be of a greatest importance (Table 14.1). Even if the infectious dose vary with the strains, the age of the patient, or other parameters, some pathogens are highly dangerous for men even at low concentrations (HAV, E. coli O157:H7, V. cholerae), whereas others have to be ingested in high concentrations to be harmful (V. parahaemolyticus) or are highly infectious but not very dangerous (norovirus). Thus, for some pathogens, a low contamination in seafood, for example, is not acceptable, based on risk-assessment models (Colwell et al., 1996). Among bacteria, the Vibrio family plays an important role in infections, waterborne or seafood diseases, especially in countries surrounded by warm marine waters. Toxigenic V. cholerae O1 and O139 are the causative agents of cholera in developing countries, whereas, non-O1 and non-O139 strains may also be pathogenic but are rarely associated with epidemics (Anonymous, 2002). V. parahaemolyticus has been recognized as a major cause of foodborne gastroenteritis in Japan and Eastern countries and has been linked to seafood consumption. In the United States, more than 700 cases of illness due to V. parahaemolyticus and associated to M. Pommepuy, D. Hervio-Heath, M. P. Caprais, M. Gourmelon, J. C. Le Saux, and F. Le Guyader de Brest, BP 70, 29280 Plouzan´e, France.

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Table 14.1. Microbial reported infection doses. Microorganism Salmonella spp. Vibrio parahaemolyticus Campylobacter jejuni Listeria monocytogenes Vibrio cholerae Escherichia coli O157:H7 Hepatitis A Norovirus a

Estimated minimum infectious dose 104 –1010 106 –109 102 –109a 102 –106 103 10–102 14 months in laboratory-based experiments with seawater, when seeded at ∼106 cells/ml (Hoff, 1989). Thus, there is the potential for long-term survival in the vicinity of fish farms (Husev˚ag et al., 1991). The pathogen has been detected in the sediment (12–43 cells/ml)

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below fish farms, several months after an outbreak of disease. In addition, it has been found in the sediments from fish farms that were not experiencing clinical disease (Enger et al., 1989). It is likely that there is a reservoir of the pathogen around farmed fish, from which infections may occur. Intraperitoneal injection and immersion of Atlantic salmon and rainbow trout with broth cultures has led to the development of clinical disease, with LD50 doses ranging from 4 × 106 to 1 × 108 cells/fish (Wiik et al., 1989). 17.3.3.6. Vibrio splendidus V. splendidus, which is associated with septicemia, is a component of the normal aquatic bacterial microflora, with survival in the marine environment of >114 days being recorded (Lopez & Angulo, 1995). The organism has been reported to be pathogenic to rainbow trout and turbot, with an LD50 of 2.2 × 104 cells and 1.2 × 104 cells, respectively (Angulo et al., 1994), and to corkwing wrasse (Jensen et al., 2003). 17.3.3.7. Vibrio vulnificus V. vulnificus, which is the cause of septicemia especially in eels (Austin & Austin, 1999), is ubiquitous in the coastal marine, and estuarine environment, where it occurs routinely in low numbers (Oliver et al., 1983). The reservoir for infection is almost certainly seawater (Høi et al.,1998; Marco-Noales et al., 2001). The organism survives in brackish water and on the surfaces of eels for 14 days (Amaro et al., 1995). V. vulnificus is capable of entering eels through the skin (Amaro et al., 1995). LD50 doses are 1, a condition that occurs as D increases. In Eq. (3), the number of prepatents increases by the number of susceptibles that become infected in Eq. (1), and decreases by the number prepatents that develop into acutes. Equation (3) models the abundance of acutes, which increases by the proportion of prepatents becoming acute, and decreases by the proportion of acutes either dying or developing into chronics. Equation (4) tracks the number of chronics, which increases as acutes develop chronic infections, and decreases as chronics die from the infection. Finally, in Eq. (5), the number of deads increases as animals die and decreases as deads become non-infectious.

19.2.2. R0 and Threshold Two important outcomes of a mathematical approach to epidemics are the parasite’s basic reproduction number (R0 ) and the host threshold density. Both are useful to an assessment of the risk of an outbreak and establishment of an infection. The parasite’s basic reproduction number is analogous to the net reproduction rate of free-living organisms. The free-living organism’s net reproduction rate equals the total number of female offspring produced by a female during her lifetime. Similarly, the parasite’s basic reproduction number equals the number of new infections transmitted from a single infected host over the duration of that host’s infectious period. It is calculated as the mean life span of an infection times the number of transmissions that would occur over that infectious period (Mollison, 1995). Simply, if R0 > 1, the number of infections will grow and an epidemic will occur, and, if R0 < 1, the number of infections will shrink and an epidemic will not occur. The net reproduction rate of free-living organisms can be calculated from a life table, which provides a schedule of age-specific (or stage-specific) mortality and fecundity rates. Similarly, a parasite’s basic reproduction number can be determined from an infection life table such as shown in Table 19.1 provides a schedule of stage-specific durations and transmissions. For our hypothetical pathogen, the mean life expectancy of each stage of infection is 1/ν for the prepatent stage, 1/(1 − (1 − αA )(1 − ρ)) for the acute stage, 1/αC for the chronic stage, and 1/δ for the dead carcass stage. All infected animals go through the prepatent, acute, and carcass stages, whereas only a portion, (ρ − αa ρ)/()1 − (1 − αa ) (1 − ρ), of acutes develop Table 19.1. Infection life table. Stage Prepatent Acute Chronic Dead

Mean duration 1 ν 1 1 − (1 − αA )(1 − ρ) 1 ρ − αa ρ αC 1 − (1 − αa ) (1 − ρ) 1 δ

Transmissions

Net transmissions

0

0

βA S βC S βD S

βA S 1 − (1 − αA )(1 − ρ) βC S ρ − αa ρ αC 1 − (1 − αa ) (1 − ρ) βD S δ

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into the chronic stage. Therefore, the life expectancy of a complete infection is 1 1 1 1 ρ − αa ρ + + + . ν 1 − (1 − αa ) (1 − ρ) αc 1 − (1 − αa ) (1 − ρ) δ

(6)

The transmission potential of each stage of the infection to transmit (analogous to fecundity) varies. For example, since the prepatent stage is not infectious, the transmission potential is zero, and the infectious period encompasses only the acute, chronic, and carcass stages. R0 is βa S0 ρ − aa ρ βc S0 βd S0 + + . 1 − (1 − αa ) (1 − ρ) δ αc 1 − (1 − αa ) (1 − ρ)

(7)

If we solve for S0 in expression (7) when R0 is equal to 1, we get the threshold number of susceptible hosts needed for an epidemic to occur, αc δ (−αa − ρ + αa ρ) . −βa δαc αa − βd αc ρ − βd αc αa + βd αc αa ρ − βc ρδ + βc ρδαa

(8)

The threshold (S0 ) and R0 values from these equations are only exact for host populations with a constant number of susceptibles, so that S does not change; however, the S0 and R0 values are useful for comparing the capacity for the spread of two pathogens. They indicate in a comparative way the relative risks associated with various pathogens in aquaculture or in the wild. Expressions (7) and (8) can give us insights into how particular parameters influence the likelihood of an epidemic. Increases in transmission probabilities (β-values) and the density of susceptibles will increase R0 (value of Eq. (7)), whereas increases in virulence, recovery, and removal rates all will negatively affect the R0 . Because R0 is a function of host density, the likelihood of an epidemic depends not only on the characteristics of the pathogen but also on the characteristics of the local host population. In addition, the value of each of the parameters of the general model will vary for each host–pathogen combination.

19.3. SHRIMP AQUACULTURE PATHOGENS The dynamic approach to the epidemiology of microparasite pathogens outlined above can be used to begin to evaluate the potential for the spread of specific microparasite pathogens in marine environments. As examples, viruses of importance to shrimp aquaculture have been chosen because the industry faces a very substantial problem resulting from viral diseases (EPA, 1999; Lightner, 1996). Clearly, pathogens threaten aquaculture industries. The dissemination of pathogens to aquaculture throughout the world is being affected by transport of live shrimp for culture (Nakano et al., 1994) and through the shipment of frozen product carrying infectious pathogens (Durand et al., 2000; Nunan et al., 1998). However, what is often less well appreciated is the threat that arises from the unintentional introduction or widespread enhancement of these viruses in wild stocks of shrimp. In areas that are presently affected by these pathogens in cultured shrimp, large amounts of virus enter native waters during routine water exchange and during emergency drain-harvests (Chantanachookin et al., 1993). If these viruses become established in wild stocks of shrimp, those stocks can become reservoirs for infecting pondreared shrimp through contaminated water exchange, stocking of infected wild post-larvae, from seed produced from infected wild spawners, or from other animals that fed on infected individuals. In addition to the negative influence on aquaculture, the viruses could be a factor

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that alone or in combination with environmental degradation might result in a reduction of wild harvest of the affected species. We will examine Taura syndrome virus (TSV), WSSV, infectious hypodermal and hematopoietic virus (IHHNV), and yellow head virus (YHV), with particular reference to the biological characteristics that will be important to the question of epidemics of these pathogens in animal populations. In addition to attempting an analysis that is consistent with what we have presented above, we will discuss other epidemiologically significant characteristics that have not yet been incorporated into the above model. We will also note what data are lacking for a complete evaluation of the risk associated with particular pathogens.

19.3.1. Taura Syndrome Virus TSV was recognized in 1994 (Brock et al., 1995) as the causative agent of Taura syndrome, a disease that had been observed in Ecuador as early as 1992. TSV is a 30–32 nm, icosahedral virus particle containing positive-sense, single-stranded RNA of about 10.2 kb in length (Mari et al., 2002). It was allied by Bonami et al. (1997) with the family Picornaviridae based on physical and chemical characteristics, a conjecture supported by Robles-Sikisaka et al. (2001). However, from nucleotide sequence data, Mari et al. (2002) suggested that TSV is more closely related to the cricket paralysis-like viruses (CrPV-like viruses) and is probably a member of the “picornavirus superfamily”, the family Dicistroviridae, and the genus Cripavirus of which it is a candidate species. TSV is one of the most important viral pathogens affecting the aquaculture of Litopenaeus vannamei, the most commonly cultured shrimp in the United States (Brock et al., 1995; Hasson et al., 1995) and may pose a risk to natural populations as well (OIE Fish Diseases Commission, 2002; Overstreet et al., 1997). Taura disease is a rapidly progressing disease marked by extensive mortalities that become evident 25–35 days after a shrimp pond is stocked with post-larval shrimp (Brock et al., 1995). Elevated death rates last only a matter of days but commonly reach 25% per day and leave a mere 5–25% of the shrimp alive. Survivors carry chronic infections at least a year and probably for life (Lotz, 1997b). 19.3.1.1. Life-Cycle Analysis Lotz et al. (2003) identify five epidemically important stages in the life cycle of TSV. Therefore, the life-cycle diagram depicted in Figure 19.2 serves as the life-cycle diagram of TSV, and Eqs. (1)–(5) represent the mathematical form of the life-cycle diagram. Lotz et al. (2003) estimated the parameter values of TSV in juvenile L. vannamei (Table 19.2). Estimates were obtained using 12 susceptible shrimp in a 1 m2 tank, for a time unit of 24 h, and an initial dose of one infected shrimp. Lotz et al. (2002) arbitrarily chose 0.1 for the removal coefficient, δ. More recent work in our laboratory has shown that 0.9 is a better estimate for the removal coefficient. Using Eqs. (7) and (8) and the values in Table 19.2, we estimate R0 to be 31 and the threshold density to be 0.39 L. vannamei/m2 . 19.3.1.2. Vertical Transmission TSV appears not to be transmitted vertically through the oocytes of the female to the nauplii. Hasson et al. (1999) reported that TSV is concentrated in the lymphoid organ of Taura

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Table 19.2. Estimates of epidemic parameters for TSV. Parameter

Estimate (95% CI)

Transmission (βA ) Transmission (βC ) Transmission (βD ) Patency (ν) Virulence (αA ) Virulence (αC ) Recovery (ρ)

0.10 (0.02–0.15) 0.10 (0.02–0.15) 0.58 (0.40–0.75) 0.49 (–) 0.40 (–)