Aquatic animals : biology, habitats, and threats 9781614701231, 1614701237

Table of contents :
AQUATIC ANIMALS
AQUATIC ANIMALS
CONTENTS
PREFACE
ACOUSTIC ECOLOGY OF PINNIPEDS IN POLAR HABITATS
ABSTRACT
INTRODUCTION
What is Acoustic Ecology?
POLAR REGIONS
SUB POLAR REGIONS
POLAR REGIONS AS HABITAT FOR PINNIPEDS
Adaptations for Life in An Ice Habitat
Different Species Use Different Ice Types
The Role of Vocalizations for Polar Pinnipeds
POLAR PINNIPED ACOUSTIC ECOLOGY
Impact of Ice on the Acoustic Environment
Seasonal Variation in the Biotic Soundscape and Inter-Specific Competition for Acoustic Space
Territoriality in an Unstable Environment
ENVIRONMENTAL THREATS
Habitat Modification
Ecosystem Alteration
Stresses to Body Condition and Health
Increased Human Interactions
CASE STUDY: ARCTIC PINNIPEDS
Acoustic Data: Bering Sea
Acoustic Activity of Arctic Pinnipeds in Relation to Ice Conditions
Inter-Specific Vocal Comparison of Arctic Pinnipeds
CASE STUDY: ANTARCTIC PINNIPEDS
Acoustic Data: PALAOA
Acoustic Activity of Antarctic Pinnipeds in Relation to Ice Conditions
CONCLUSION
What Can We Learn by Taking a Bi-Polar/Comparative Approach?
Outlook
ACKNOWLEDGMENTS
Reviewed by
REFERENCES
IMMUNOTOXICOLOGICAL REACTIVITY OF HEMOCYTE OF JUVENILE MUDCRAB OF SUNDARBANS BIOSPHERE RESERVE OF INDIA
ABSTRACT
INTRODUCTION
MATERIALS AND METHODS
1. Collection, Transportation and Maintenance of S. Serrata
2. Treatment with Sodium Arsenite
3. Collection of Hemolymph
4. Testing of Cell Viability
5. Determination of Total Hemocyte Count (THC)
6. Aggregation Assay
7. Non-Self Surface Adhesion Assay
8. Assay of Phagocytic Response Under the Challenged Yeast
9. Activity of Superoxide Anion (O2-)
10. Activity of Nitric Oxide (NO)
11. Activity of Phenoloxidase (PO)
12. Estimation of Protein
13. Statistical Analyses
RESULT AND DISCUSSION
ACKNOWLEDGMENT
REFERENCES
THE POTENTIAL THREAT OF GENOTOXIC METALS TO MARINE MAMMAL HEALTH: A CASE STUDY OF CHROMIUM TOXICITY IN TOOTHED AND BALEEN WHALES
ABSTRACT
INTRODUCTION
MATERIALS AND METHODS
Biopsy Collection
Development of Whale Cell Lines
Cr(VI) Compounds
Inductively Coupled Plasma Mass Spectroscopy
Measuring Cell Death
Measuring Clastogenicity
RESULTS
Determining Cr in Levels in Whale Skin
Evaluating the Toxicity of Cr to Cetaceans
Toxicity Context from the Cr Literature
Toxicity Context from Whale Cell Culture Studies
Cr Cytotoxicity Studies in Whale Cells
Cr Genotoxicity Studies in Whale Cells
CONCLUSION
REFERENCES
HYALELLA GENUS: ANTHROPOGENIC THREATS TO A BIOINDICATOR
ABSTRACT
INTRODUCTION
OBJECTIVE
Hyalella Genus and Chemical Stressors
Biomarkers as Means of Assessing the Toxicity
CONCLUSION
REFERENCES
THE AROMATIC HYDROCARBON RECEPTOR MEDIATED CYTOCHROME P450 1A INDUCTION IN AQUATIC ANIMALS: BIOMONITORING OF ORGANIC POLLUTION IN AN AQUATIC ENVIRONMENT
ABSTRACT
1. INTRODUCTION
2. CYTOCHROME P450 – ORIGIN AND EVOLUTION
3. MECHANISM OF CYTOCHROMEP 450
4. CYTOCHROME P450 IN AQUATIC ORGANISMS
5. CATALYTIC CYCLE OF CYP450 AND MONOOXYGENASE REACTION
6. CYP450 MEDIATED XENOBIOTIC (BENZO(A)PYRENE) METABOLISM IN AQUATIC ORGANISMS
7. AROMATIC HYDROCARBON RECEPTOR (AHR) AND ITS MECHANISM
8. AHR IN AQUATIC ORGANISMS
9. CYP1A AS A BIOMARKER TO ENVIRONMENTAL POLLUTION
10. FUTURE DIRECTIONS
ACKNOWLEDGMENTS
REFERENCES
ROLE OF BACTERIA IN THE CHILLED STORAGE AND CRYOPRESERVATION OF SPERM IN AQUATIC ANIMALS: A REVIEW
ABSTRACT
1. A SHORT OVERVIEW OF THEORY ON CHILLED STORED- AND CRYOPRESERVED SPERM
1.1. Sperm Quality
1.2. Short-Term Storage of Sperm
1.3. Long-Term Storage of Sperm
2. BACTERIAL CONTAMINATION IN THE CHILLED STORED- AND CRYOPRESERVED SPERM
2.1. Occurrence of Bacteria in Aquatic Animals
2.2. Type and Number of Bacteria in Sperms of Aquatic Animals
2.3. Bacterial Profiles in the Chilled Stored Sperm/ Spermatophores of Aquatic Animals
2.3.1. Effect of Extenders on Bacterial Profiles in the Chilled Storage Process
2.3.2. Role of Contaminated Bacteria in Chilled Storage Process
2.3.3. Effect of Antibiotics on Contaminated Bacteria in Chilled Stored Sperm
2.4. Bacterial Profiles in Cryopreserved Sperm/ Spermatophores of Aquatic Animals
2.4.1. Change in Bacterial Flora of Cryopreserved Sperm/Spermatophores During Storage in Liquid Nitrogen
3. BIOSECURITY AND RISK ASSESSMENT OF CRYOPRESERVED SPERM
4. CONCLUDING REMARKS AND RECOMMENDATIONS
ACKNOWLEDGMENTS
REFERENCES
INDEX

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MARINE BIOLOGY

AQUATIC ANIMALS BIOLOGY, HABITATS AND THREATS No part of this digital document may be reproduced, stored in a retrieval system or transmitted in any form or by any means. The publisher has taken reasonable care in the preparation of this digital document, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained herein. This digital document is sold with the clear understanding that the publisher is not engaged in rendering legal, medical or any other professional services.

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Library of Congress Cataloging-in-Publication Data Aquatic animals : biology, habitats, and threats / editor, David L. Eder. p. cm. Includes index. ISBN 978-1-61470-131-6 (eBook) 1. Aquatic animals--Microbiology. 2. Aquatic animals--Toxicology. 3. Aquatic animals--Diseases. 4. Aquatic habitats. I. Eder, David L. QR106.A68 2011 591.76--dc23 2011019863

Published by Nova Science Publishers, Inc. † New York

MARINE BIOLOGY

AQUATIC ANIMALS BIOLOGY, HABITATS AND THREATS

DAVID L. EDER EDITOR

Nova Science Publishers, Inc. New York

MARINE BIOLOGY Additional books in this series can be found on Nova‘s website under the Series tab.

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CONTENTS Preface

vii

Chapter 1

Acoustic Ecology of Pinnipeds in Polar Habitats Ilse C. Van Opzeeland and Jennifer L. Miksis-Olds

1

Chapter 2

Immunotoxicological Reactivity of Hemocyte of Juvenile Mudcrab of Sundarbans Biosphere Reserve of India 53 Sajal Ray, Sanjib Saha and Mitali Ray

Chapter 3

The Potential Threat of Genotoxic Metals to Marine Mammal Health: A Case Study of Chromium Toxicity in Toothed and Baleen Whales Carolyne LaCerte and John Pierce Wise, Sr.

77

Chapter 4

Hyalella Genus: Anthropogenic Threats to a Bioindicator 99 Guendalina Turcato Oliveira, Felipe Amorim Fernandes and Bibiana Kaiser Dutra

Chapter 5

The Aromatic Hydrocarbon Receptor Mediated Cytochrome P450 1a Induction in Aquatic Animals: Biomonitoring of Organic Pollution in an Aquatic Environment S. Arun

Chapter 6

Index

Role of Bacteria in the Chilled Storage and Cryopreservation of Sperm in Aquatic Animals: A Review Subuntith Nimrat and Verapong Vuthiphandchai

113

139 191

PREFACE This book presents topical research from across the globe in the study of the biology, habitats and threats to aquatic animals. Topics discussed include the acoustic ecology of pinnipeds in polar habitats; the immunotoxicological reactivity of hemocytes of juvenile mudcrabs; the potential threat of genotoxic metals to marine mammal health; toxic contaminants in aquatic medium concerns and the role of bacteria in the chilled storage and cryopreservation of sperm in aquatic animals. Chapter 1 - The Arctic and Antarctic share a strong seasonality of light and temperature extremes that create similar selective pressures for animals living in these two polar and associated sub polar regions. There is significant evidence that Arctic ice conditions are rapidly deteriorating due to global climate change and evidence that Antarctic ice is also changing. Sub polar regions experience a seasonal ice cover with an added dynamic range of environmental conditions often exceeding that at the poles. In polar and sub polar waters, the presence of sea ice dominates the physical environment for an extended period of time. This strongly influences pinniped distribution, reproductive strategies, foraging ecology, and acoustic behavior. In both polar regions, seals have distinctly different underwater vocal repertoires associated with breeding and an airborne repertoire associated with mother and pup communication. In this chapter, a comparative approach is taken to relate the underwater acoustic behavior of polar pinnipeds (phocids and walrus) to their ecology and aspects of their sea ice habitat. Understanding the commonalities and differences in the spatial, spectral, and temporal characteristics of vocalizations from species with comparable biologies relative to local sea ice conditions may provide insights into the acoustic ecology of pinnipeds in polar habitats.

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David L. Eder

Acoustic ecology describes the interaction between an animal and its environment as mediated through sound and is determined by the species' behavioral ecology, but also by biotic and abiotic factors of the environment. Insight into acoustic ecology may therefore provide information on the potential direct and indirect impacts of polar climate change on pinnipeds. Loss of sea ice will likely affect polar pinniped distribution and breeding activities; both of which can be detected through long-term passive acoustic monitoring. One of the secondary effects of sea ice loss is the impact to the polar acoustic soundscape (i.e., the acoustic environment). The communication system of polar pinnipeds, which permeates critical life functions like breeding, evolved in an acoustic environment dominated by ambient noise levels associated with seasonal ice cover. Changes in ice dynamics will likely be accompanied by shifts in habitat use by both marine animals and humans which will lead to a corresponding change in the acoustic soundscape. The combined impacts of these effects is unprecedented, but the knowledge gained from a comparison of pinniped acoustic ecology between the Arctic and the Antarctic will lead to a better understanding of the relationship between ice seals and their changing environment. Chapter 2 - Sundarbans Biosphere Reserve of India supports a wide range of flora and fauna. Ecologically, this reserve is a mangrove forest with numerous estuarine creeks. Mudcrab Scylla serrata (Crustacea: Decapoda) is an important member of estuarine habitat of Sundarbans. Indiscriminate collection of juvenile crab seeds and contamination of estuarine water pose a serious threat for this aquatic animal. Generally invertebrates including mudcrab lack adaptive immune system and depend on innate immune defense to combat pathogen and toxin . Cellular immune response involved the recruitment of immunocompetent hemocytes capable of performing adhesion, aggregation, phagocytosis, generation of intracellular cytotoxic molecules etc. In this chapter, spectrum of innate immunological reactivity was examined in juvenile mudcrab under the exposure of arsenic, a toxic metalloid of aquatic environment. Experimental juveniles were exposed to 1, 2 and 3 ppm of sodium arsenite for 1, 2, 3 and 4 days in controlled laboratory condition. Total hemocyte count is a stress indicator of pollution which was enumerated both in the control and arsenic exposed S. serrata. Crabs exposed to sodium arsenite expressed an alteration in total hemocyte density. Aggregation and adhesion of hemocytes are metabolic behaviours whereas phagocytosis is a classical immunological response of hemocytes and arsenic exposure yields a dose responsive decrease in aggregation, adhesion and phagocytosis against the

Preface

ix

control indicating perturbation of blood cell hemeostasis, alteration of hemocyte surface characteristics and immune suppression. Capability of generation of cytotoxic molecules like superoxide anion and nitric oxide are reported as a potential strategy of immunological defense against pathogen invasion. Disruption of hemocyte density, adhesion, aggregation, phagocytosis and increment of cytotoxic response may lead to a state of alteration of immunity. Arsenic induced alteration of immune status may impart a state of vulnerability in juvenile crab inhabiting the toxic aquatic environment. Such a situation may reduce the ecological fitness of mudcrab to combat against the toxic challenge caused by arsenic pollution. Chapter 3 - Ocean pollution has emerged as one of the greatest threats to marine animal heath with recent data showing that metal pollution has reached even the remotest ocean regions. Marine mammals are at particular risk as they integrate all possible routes of exposure: inhalation, ingestion and dermal pathways. Many marine mammals have become critically endangered and are failing to recover. Marine mammals are important species for ecosystem health and for the economies of coastal communities. Moreover, they are charismatic species that capture people's attention and remind them of the importance of the ocean environment. Understanding them and the threats that face them is a key need in protecting the ocean environment. The authors have been pioneering the study of metals, particularly genotoxic metals as particular threat to marine mammal health. Marine mammals are exposed to metals and if these metals can damage DNA in these animals, the consequences could impact both the individual, through metalinduced disease, and the population as a whole, by impairing reproduction. In this chapter, the authors focus on chromium as a representative genotoxic metal and describe the levels of Cr in tissues from free ranging, healthy whales considering both toothed and baleen whales. They then determine the genotoxic effects of Cr in cultured whale cells and compare genotoxic doses to tissue levels. The chapter discusses the potential threat of genotoxic metals, like chromium, what the sources of exposure are in the marine environment, and compares the data to effects seen in humans and laboratory animals. Chapter 4 - In this chapter the authors review the current knowledge about the usage of Hyalella genus as a bioindicator using reproductive and biochemical markers to assess damage, as well as threats (pesticides, heavy metals and endocrine disruptors) that could lead the significant alteration of the physiological pattern of these animals. Biochemical markers have the advantage of demonstrating a rapid response from animal to stress, while the

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David L. Eder

reproductive show a high biological and ecological significance, because they demonstrate long-term effects that may lead to population extinction, and indicate the degree of sensitivity. Hyalella are common genus in the Americas, with 51 described species found in a variety of freshwater habitats, often cling to the vegetation, swim in the water, or burrow in the sediment where its constitute important links in the food web, serving to transfer energy from basic recourse (detritus and algae) to higher-level consumers. The structure of natural communities, the richness and diversity of species, and the interactions among them have profound implications for ecosystem functioning, modulating the fluxes of energy and materials within and among environments. The human activities have directly or indirectly added novel stressors to natural aquatic ecosystems which have the potential for affecting the structure and function of natural communities and among this stressor we can list the pesticides, the heavy metals, and home and industrial sewage that have been reported to cause dramatic changes in freshwater environments. Recently, a great deal of attention has been devoted to the use of physiological and energetic processes of non-target organisms as sensitive indicators of toxic stress. We discuss the current information on these crustaceans amphipods as models for ecotoxicology and describe areas of future research. Chapter 5 - There is a serious concern about the disposal of hazardous substances in to the aquatic medium. Toxic contaminants such as Polycyclic Aromatic Hydrocarbons (PAHs) and Polychlorinated Biphenyls (PCBs) are ubiquitous global contaminants that affect both the quality of water and the hygiene of aquatic organisms. Toxicity mediated through these contaminants occurs through the cytoplasmic protein called Aromatic Hydrocarbon Receptor (AhR) which is bound to two other proteins: HSP90 and XAP2. Upon binding with the contaminant AhR translocates to the nucleus, where they dissociate from HSP90 and XAP2, and forms a heterodimer complex with bHLH-PAS and the AHR nuclear translocator (ARNT). Together, they bind to Xenobiotic (contaminants / Aromatic hydrocarbons) Responsive Elements (XRE) to regulate the expression/ induction of Cytochrome P4501A (CYP1A). Thus, the induction of CYP1A through the activation of AhR has been used as a biomarker for exposure to organic pollutants in various aquatic organisms ranging from invertebrates to vertebrates. CYP1A activity has been most extensively analyzed using the ethoxyresorufin O-deethylase (EROD) or aryl hydrocarbon hydroxylase (AHH) assays. In this chapter, the author describes the molecular mechanism of AhR mediated CYP1A induction in aquatic organisms and how it could be applied as a suitable biomarker for the early detection of organic pollution in an aquatic environment.

Preface

xi

Chapter 6 - Bacterial contamination is considered a major issue for the management of sperm preservation in aquatic animals. Chilled storage and cryopreservation of aquatic animal sperm have been used to meet the growing demand for seeds. Chilled storage of sperm has been effectively used in a number of fish, but in marine invertebrates has been limited to shrimp, lobster, echiura and shellfish. Successful and reproducible cryopreservation has proven to be a reliable method for preserving genetic lines of fish for aquaculture and conservation, but among invertebrates is limited to polycheates, cnidaria, crustaceans (shrimp and crab), shellfish (oysters, hard clams and abalone) and echinoderms (sea urchin and sea cucumber). Few studies have documented changes in bacterial type and number during chilled storage or cryostorage. Access of good quality chilled-stored or cryopreserved sperm is necessary for the management of risk associated with sperm preservation. The presence of bacteria in sperm of aquatic animals can result in deleterious effects on sperm quality and viability if left uncontrolled. Potential pathogens present on the preserved sperm of aquatic animals may affect the health of broodstock and offspring. This paper summarizes the current knowledge pertaining to bacterial contamination of aquatic animal sperm, including bacterial types and sources of contaminants, and gives an overview of the incidence of bacterial contamination, biosecurity and risk assessment and effects of bacterial contamination on offspring quality. These ideas, if taken into account in commercial aquaculture and conservation of aquatic animals, are almost certain to increase the value added or protection of at least some cultured species. Microbial contamination risk of chilled-stored or cryopreserved sperm after short-term or long-term storage is discussed. Future direction and new avenues of future research on chilled storage and cryopreservation of sperm in aquatic animals are proposed.

In: Aquatic Animals Editor: David L. Eder

ISBN: 978-1-61470-123-1 © 2012 Nova Science Publishers, Inc.

Chapter 1

ACOUSTIC ECOLOGY OF PINNIPEDS IN POLAR HABITATS Ilse C. Van Opzeeland1 and Jennifer L. Miksis-Olds2 1

2

Alfred Wegener Institute, Bremerhaven, Germany Applied Research Laboratory, The Pennsylvania State University, State College, PA, U.S.

ABSTRACT The Arctic and Antarctic share a strong seasonality of light and temperature extremes that create similar selective pressures for animals living in these two polar and associated sub polar regions. There is significant evidence that Arctic ice conditions are rapidly deteriorating due to global climate change and evidence that Antarctic ice is also changing. Sub polar regions experience a seasonal ice cover with an added dynamic range of environmental conditions often exceeding that at the poles. In polar and sub polar waters, the presence of sea ice dominates the physical environment for an extended period of time. This strongly influences pinniped distribution, reproductive strategies, foraging ecology, and acoustic behavior. In both polar regions, seals have distinctly different underwater vocal repertoires associated with breeding and an airborne repertoire associated with mother and pup communication. In this chapter, a comparative approach is taken to relate the underwater acoustic behavior of polar pinnipeds (phocids and walrus) to their ecology and aspects of their sea ice habitat. Understanding the commonalities and differences in the spatial,

2

Ilse C. Van Opzeeland and Jennifer L. Miksis-Olds spectral, and temporal characteristics of vocalizations from species with comparable biologies relative to local sea ice conditions may provide insights into the acoustic ecology of pinnipeds in polar habitats. Acoustic ecology describes the interaction between an animal and its environment as mediated through sound and is determined by the species' behavioral ecology, but also by biotic and abiotic factors of the environment. Insight into acoustic ecology may therefore provide information on the potential direct and indirect impacts of polar climate change on pinnipeds. Loss of sea ice will likely affect polar pinniped distribution and breeding activities; both of which can be detected through long-term passive acoustic monitoring. One of the secondary effects of sea ice loss is the impact to the polar acoustic soundscape (i.e., the acoustic environment). The communication system of polar pinnipeds, which permeates critical life functions like breeding, evolved in an acoustic environment dominated by ambient noise levels associated with seasonal ice cover. Changes in ice dynamics will likely be accompanied by shifts in habitat use by both marine animals and humans which will lead to a corresponding change in the acoustic soundscape. The combined impacts of these effects is unprecedented, but the knowledge gained from a comparison of pinniped acoustic ecology between the Arctic and the Antarctic will lead to a better understanding of the relationship between ice seals and their changing environment.

INTRODUCTION What is Acoustic Ecology? Marine mammals produce and use sound in many behavioral contexts, such as foraging, social communication, and orientation. The sounds that an animal produces are shaped by abiotic and biotic factors from its environment and its behavioral ecology. Figure 1 represents the various factors and interactions and how these can influence the acoustic behavior of polar marine mammals1. Together, these factors form the acoustic ecology of a species. Acoustic ecology describes the interaction between an animal and its environment as mediated through sound (e.g., Clark et al., 2009). In the following paragraphs of this section, each of the factors and interactions represented in Figure 1 is further explained. 1 Although the figure presents examples of abiotic factors specific to marine mammals in polar oceans (e.g.,ice conditions), the basic figure in principle applies to all animals relying on sound for critical aspects of their behavior.

Acoustic Ecology of Pinnipeds in Polar Habitats

3

Behavioral ecology is the evolutionary and ecological basis for animal behavior and the role of behavior in the adaptation of an animal to its living environment (Krebs & Davies, 1993). Behavioral ecological factors that can influence vocal behavior (Figure 1, arrow A) are mating strategies (e.g., territorial, roaming/non-territorial), distribution patterns (e.g., foraging behavior, migratory behavior, and home range) and reproductive characteristics (e.g., gestation, lacation patterns and duration, and time of estrus). For example, the local distribution pattern of a species will determine the range over which an animal communicates with conspecifics and thereby the acoustic characteristics of communication signals. The behavioral ecology of an animal also is influenced by abiotic and biotic factors from the environment (Figure 1, arrows B and C). The presence of prey and predators can for example affect marine mammal distribution patterns (e.g., Piatt & Methven, 1992, Figure 1, arrow C), whereas ice conditions influence mating strategy (Figure 1, arrow B, e.g., Van Parijs, et al., 2004). Biotic factors can influence acoustic behavior under certain conditions, such as the presence of predators that can lead to reduced vocal activity of prey (e.g., in the case of marine mammals hunted by killer whales, Orcinus orca, Thomas et al., 1987; Jefferson et al., 1991) or when specific feeding calls are produced during foraging (e.g., humpback whales, Megaptera novaeangliae, D‘Vincent et al., 1985, Figure 1, arrow D). Abiotic factors can influence vocal behavior directly (Figure 1, arrow E). Diel ambient temperature variations affect the timing of vocal periods in lions (Panthera leo), elephants (Loxodonta africana africana), coyotes (Canis latrans), and wolves (Canis lupus) so signallers can maximize their chance of being heard over the longest possible distances (Larom et al., 1997). Soundscape can be considered a purely abiotic factor because the physical measurement of sound levels does not separate contributing sources. However, given its importance and the many specific interactions with other acoustic ecological factors, it is shown in a separate box in Figure 1. The natural soundscape is influenced by many elements that are of biotic (e.g., sounds produced by conspecifics and prey, arrows S1, S2, respectively) and abiotic origin (e.g., glacier calving, iceberg collisions, precipitation, waves, arrow S3). Furthermore, the local soundscape in turn can affect acoustic behavior and biotic factors (indicated by the double arrows S1 and S2). Changes in local soundscape can result in changes in the source level of animal vocalizations (Brumm & Slabbekoorn, 2005). Increase in source level in response to increases in background sound level is known as the ―Lombard effect‖ (Lombard, 1911; Lane & Tranel, 1971; Pick et al., 1989) and has

4

Ilse C. Van Opzeeland and Jennifer L. Miksis-Olds

been documented in beluga (Scheifele et al., 2005) and North Atlantic right whales (Parks et al., 2011). Abiotic factors also influence biotic factors which can ultimately affect calling behavior, e.g., when ice conditions drive prey distribution, thereby indirectly determining the timing of vocal activity of marine mammals (e.g., the production of specific feeding calls, Figure 1, arrow F). Anthropogenic factors such as climate-driven changes can also impact ice conditions and prey distribution, thereby having the potential to indirectly influence acoustic behavior (Figure 1, arrows G and H, respectively). Human-generated underwater noise can impact local soundscapes and thereby affect acoustic behavior and biotic factors (Figure 1, arrow S4). Finally, anthropogenic factors can also directly influence the behavioral ecology of a species by affecting behavior or distribution patterns indicative of non-lethal disturbance effects of human activity (e.g., Frid & Dill, 2002; Branch et al., 2007; Wirsing et al., 2008, Figure 1, arrow I). The concept of acoustic ecology is a framework to explain how various factors shape vocal behavior and function as a driver determining the relation between physical presence and the probability of acoustic detection (Figure 1). Acoustic ecology provides a structure for predicting the temporal and spatial scales over which acoustic presence can be detected. The abiotic and biotic characteristics of a species‘ living environment and its behavioral ecology ultimately determine the acoustic features of its vocal behavior, the effective range of communication between conspecifics, the time of day to be vocal, seasonality of vocal behavior, and whether to vocalize or remain silent. Acoustic data is in many cases almost per definition ‗presence-only‘ data. Nevertheless, once the understanding of the acoustic ecology of a species is sufficient, it is also possible to draw conclusions on absence of animals based on acoustic data (Figure 1; Van Opzeeland, 2010). An understanding of acoustic ecology is therefore of importance to enable interpretation of acoustic data in a regionally and seasonally appropriate context (Van Parijs et al., 2009). In this chapter, the vocal behavior of pinnipeds in Arctic and Antarctic polar habitats are described and compared. Two case studies of Arctic and Antarctic pinniped species illustrate how bipolar comparisons can promote the understanding of the commonalities and differences in the spatial, spectral, and temporal characteristics of vocalizations from species with comparable biologies relative to local sea ice conditions. Even though they evolved separately, they developed remarkably similar acoustic adaptations.

Acoustic Ecology of Pinnipeds in Polar Habitats

5

Acoustic observations

Behavioral ecology Mating system Distribution

Physical presence

Acoustic presence

Acoustic behavior Call rates Temporal variability

S1

Abiotic factors

Biotic factors

Ice conditions Water/air temperature

Prey Predators

S3

S2

Soundscape

S4

I

Anthropogenic factors Climate change Underwater noise

Figure 1. Schematic representation of the acoustic ecology of polar pinnipeds, comprising the interactions between the natural factors of behavioral ecology, biotic and abiotic factors relating to acoustic behavior, (blue boxes with thin blue arrows) and the potential influences of anthropogenic factors (yellow box with thin yellow arrows). Soundscape (small blue box) is in principle considered an abiotic factor, but given the many interactions with other factors (thin arrows S1-S4), it is presented separately. The dashed line between the yellow and blue boxes separates natural factors and interactions from anthropogenic influences within the schematic representation of acoustic ecology. Green boxes and thick round arrows represent the interrelation between physical presence of animals and the probability of acoustic detection (acoustic presence), which is mediated by their acoustic ecology (Van Opzeeland, 2010).

Detailed knowledge of the acoustic ecology of polar pinnipeds is likely to contribute to a better understanding of the mechanisms by which

6

Ilse C. Van Opzeeland and Jennifer L. Miksis-Olds

anthropogenic factors, such as climate change and anthropogenic noise, impact various aspects of their life functions and needs. The characteristics of polar habitats and how these parameters contribute to the acoustic ecology of polar pinnipeds are discussed in the following section.

POLAR REGIONS The earth‘s polar oceans comprise the Southern Ocean, with the Antarctic Convergence being recognized as its natural northern boundary (CCMLAR, 1980), and the Arctic Ocean, i.e., the ocean located within the Arctic Circle occupying the region around the North Pole (Figure 2).

Maps from Van Opzeeland (2010). Figure 2. Maps of the earth‘s polar oceans. Left: the Southern Ocean, with the Antarctic Convergence (white dotted line, schematic representation of mean position of Antarctic Convergence) as its northern boundary; right: the Arctic Ocean and the Arctic Circle (white dotted line).

The Arctic Ocean is largely land-locked and buffered by extensive shallow shelf seas which are influenced by seasonal air and freshwater fluxes from the surrounding continents. The Arctic Ocean has a permanent cover of slowly circulating multi-year ice floes surrounded by a zone of seasonal pack ice and a zone of land-fast ice (Stonehouse, 1989; Figure 3). In winter, sea ice cover can reach a maximum of 13.9x106 km2, while in summer approximately 6.2x106 km2 of the Arctic Ocean can be ice-covered (Johannessen et al., 1999). Sea ice coverage in the far North is relatively

Acoustic Ecology of Pinnipeds in Polar Habitats

7

stable, with sea ice melt occurring primarily at the periphery. Consequently, a large part of the Arctic sea ice consists of multi-year ice seasonally or year round. However, this situation may change rapidly within the next decades, as the Arctic sea ice recently showed substantial decreases in both extent and thickness in response to global warming (IPCC, 2007).

Boreal summer

Boreal winter

Maps provided by www.seaice.de which are based on algorithms described in Spreen et al. (2008). Figure 3. Maps of the sea ice concentration in the Arctic. Left: Example of a day with minimal sea ice extent (1 September 2009); Right: Example of a day with maximal sea ice extent (16 March 2009).

The Antarctic consists of a continent forming ~10% of the earth‘s land surface. The continent is surrounded by a dynamic, open ocean and an unusually deep continental shelf, a side-effect of the weight of the ice sheet covering the continent (Knox, 2007). Apart from the northern part of the Antarctic Peninsula, most of the Antarctic continent lies south of the Antarctic Circle (66° 33‘44 S), south of which continuous daylight prevails during the austral summer and darkness dominates the austral winter, with

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Ilse C. Van Opzeeland and Jennifer L. Miksis-Olds

the coastal regions experiencing twilight periods throughout winter. The natural northern boundary of the Southern Ocean is formed by the Antarctic Convergence (or Polar Front), which forms a sharp temperature boundary between northern temperate waters and southern polar waters (Figure 4). The Southern Ocean‘s sea ice canopy offers substantial seasonal habitat heterogeneity, reaching up to 20x106 km2 (effectively doubling the frozen area of Antarctica) during winter and receding to less than 4x106 km2 in austral summer (Knox 2007; Thomas et al., 2008, Figure 4).

Austral summer

Austral winter

Maps provided by www.seaice.de which are based on algorithms described in Spreen et al. (2008). Figure 4. Maps of the sea ice concentration in the Southern Ocean. Left: Example of a day with minimal sea ice extent (31 January 2009); Right: Example of a day with maximal sea ice extent (11 August 2009).

The Arctic and Antarctic share many characteristics relating to light cycles, temperature extremes, biological productivity, and sea ice distribution that create similar selective pressures for animals living in the two polar and associated sub polar regions. There are, however, striking differences that explain why for example there are no land predators, such as polar bears (Ursus maritimus) or Arctic fox (Alopex lagopus) in the Antarctic and no penguins in the Arctic, and these differences must be considered when making bipolar comparisons between similar species (summarized in

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Dieckmann & Hellmer (2010)). First and foremost, the Arctic Ocean is surrounded by landmasses with the polar ocean in the middle over the North Pole with only one deep water passage through which water is exchanged with the global oceans. Therefore, the ice formed in the Arctic is geographically constrained resulting in thick ridges as ice floes build up and converge. The Southern Ocean surrounds the Antarctic continent and effectively separates it from the rest of the world. Here the ice has the freedom to expand northward without physical constraint so ice ridges are not as common in the Antarctic, and most of the ice that forms during the winter melts in the summer (Thomas & Dieckmann 2003; 2010). In the Arctic, the presence of ridge ice creates thicker seasonal ice that has a longer life cycle which leads to ice that stays frozen longer during the summer melt. Some Arctic sea ice remains throughout the summer and contributes to multiyear sea ice, which is uncommon in the Antarctic (Thomas & Dieckmann 2003; 2010). A second major difference between the polar regions that deserves consideration is air temperature and associated climate patterns. In this chapter, the discussion is restricted to coastal and marine areas because polar pinnipeds are not encountered in inland areas. Arctic coastal regions experience a peak in summer temperatures that can be 15-30o C above monthly averages for the winter months (Stonehouse, 1989). In contrast, coastal areas of Antarctica have a less dynamic temperature range with summer monthly averages being no more than 10o C warmer than the winter averages. The greater ice and snow cover in the Antarctic compared to the Arctic results in an overall higher albedo and lower heat absorption in the Antarctic, which leads to a more uniform coastal temperature. In contrast to air temperature, water temperature above 200 m is less variable between seasons. Maximum sea surface variability for the open polar oceans is less than 4o across seasons (Comiso 2000; 2002). Seasonal changes in the temperature of the ocean at any location are seldom observed below 200 m (Knauss, 1997). In the Antarctic, Weddell seals (Leptonychotes weddellii) are seldomly seen hauled out on the ice during winter, presumably because the wind chill is too extreme (Rouget et al., 2007). Ringed seals (Phoca hispida) protect themselves from wind chill when hauled out by building and occupying subnivean lairs. The temperature of unoccupied lairs can be up to 10o warmer than ambient air temperatures, and occupied lairs can warm up to 20o more than ambient temperatures from the seal body heat (Kelly, 2005).

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The Arctic can be split into two general climatic areas: the central maritime basin and the peripheral coastal areas (Thomas et al., 2008). The central basin is characterized as a climatically stable area with a strong central anticyclone, clear skies, and light winds in winter; the anticyclone weakens in summer and brings with it moist air and stronger winds (Thomas et al., 2008). Climates of the coastal areas are more diverse depending on location. The Hudson Strait area experiences the warmest average temperatures, but heaviest snowfall and wind speeds, whereas an offshoot of the North Atlantic Drift warms the southern portion of Greenland resulting in a mild climate. Climate of the Antarctic marine system is less complicated. The relatively warmer ocean, compared to land conditions, creates a milder climate with generally uniform climatic conditions in coastal areas. Cyclonic activity dominates offshore areas, and cold katabatic winds originating from inland topography strongly influences the local weather near islands and coastal areas.

SUB POLAR REGIONS Sub polar regions are tightly coupled to the ecosystem and oceanographic dynamics of the Arctic and Antarctic regions and are included in our working definition of ―polar‖. The sub-Antarctic is loosely (depending on climate) defined herein as the area north of the Polar Front extending to approximately 45o S and includes parts of the Atlantic, Pacific, and Indian Oceans. The sub-Antarctic is characterized by glaciers found on islands (e.g., Heard, Bouvet, and Kerguelen Islands). There is no natural sub polar Arctic boundary equivalent to the Antarctic due to the amount of land surrounding the northern polar region, so the sub-Arctic is defined as the region from approximately 50o N to the Arctic Circle including the Bering Sea, Sea of Okhotsk, North Sea, Labrador Sea, and Hudson Bay. Sub polar regions are typically ice-free for a good portion of the year and have a milder climate without the extremes of the polar regions. Monthly temperatures are above 10 o C for at least one and at most three months of the year. Ice is the main feature defining the polar and sub polar regions, and it is also the major difference between the two regions. Sea ice dynamics influence regional oceanic circulation patterns, pelagic and benthic production, nutrient availability, and distribution of upper trophic level species. Winters with extensive ice cover and late ice retreats in the sub polar marine ecosystem are similar to those of polar regions; extensive seasonal sea

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ice leads to a higher water column and sediment carbon production, low grazing rates by zooplankton, and tight pelagic-benthic coupling of organic production (Grebmeier et al., 2006). Bio-physical interactions relating to the extent of ice play an important role in respect to water column stability and location of macronutrients and associated productivity. Early pelagic spring blooms that follow the ice edge retreat occur in the cold ( 60%. Compared to species in temperate regions, ice-breeding pinnipeds need to cope with much more variation in their breeding habitat. Variation exists not only among years, but also within the breeding seasons, as gradual changes in ice conditions may affect the location of under-ice features which determine territory density, size and shape.

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ENVIRONMENTAL THREATS While global climate change is predicted to increase oceanic mixing, wave generation, and coastal flooding, the reduction in sea ice is the most severe threat to polar regions and ice dependent animals (Walsh, 2008; Moore & Huntington, 2008). In addition, polar waters are important as feeding areas for many seasonally present marine mammal species, causing the effect of human-induced climate and ice changes to affect marine mammals beyond the period they spend in the polar oceans. Changes in climate are predicted to be greater and more rapid in polar regions, branding the polar ecosystems as more sensitive and reliable indicators of change than other regions of the world (Thomas et al., 2008). This chapter does not attempt to provide a detailed discussion on all the complex drivers and potential ramifications of climate change (see Arctic Climate Impact Assessment (2005) for a comprehensive review), but merely provides a context in which to consider changes to the acoustic environment of polar pinnipeds as their environment undergoes a relatively rapid transformation. The Arctic is projected to experience ―ice-free‖ summers within 30 years (Stroeve et al., 2007; Wang & Overland, 2009), making it more similar to the annual formation and melting of winter ice in the Antarctic. Leading up to this pivotal point will be drastic reductions in the extent, thickness and stability of winter ice in both the Arctic and sub-Arctic. The effects of global climate change on Antarctic regions are less obvious than those in the Arctic, with the exception of the Antarctic Peninsula (e.g., Walther et al., 2002; Dieckmann & Hellmer, 2010). In the Antarctic Peninsula (the most northerly part of the continent), the sea ice extent has been decreasing, whereas in the eastern Ross Sea region it has increased (Jacobs, 2006). The expansion in sea ice persistence and the occurrence of more concentrated sea ice fields during spring in the Ross Sea region causes the sea ice to take more time to recede in summer (Stammerjohn et al., 2008). This situation potentially affects the availability and timing of availability of specific ice types used by the different ice breeding pinniped species (Siniff et al., 2008). The decrease in persistence and extent of sea ice in other areas of the Antarctic (e.g., Bellingshausen and southern Scotia Seas), causes a different threat to ice-breeding seals inhabiting those regions. Reduction in sea ice extent may negatively impact food supply for species that feed on krill, such as crabeater and leopard seals. Antarctic krill (Euphasia superba) are highly dependent on the presence of sea ice with under-ice biota providing protection from predators and a

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zooplankton food source during austral winter and spring (Knox, 2007). Local abundance of krill is therefore likely to decrease with the sea ice. With an understanding of the observed differences between the Antarctic Peninsula and the rest of Antarctica, the consequences of sea ice reduction from the perspective of polar pinnipeds can be considered to parallel those in the Arctic. Moore and Huntington (2008) identified four general areas of challenges that Arctic marine mammals (and by extension here, Antarctic marine mammals, especially in the Antarctic Peninsula) will face as a result of changes to ice conditions: habitat modification (Laidre et al., 2008), ecosystem alteration (Bluhm & Gradinger, 2008; Fisbach et al., 2007; Stirling & Parkinson, 2006); stresses to maintain body condition and health (Burek et al, 2008), and increased human interactions (Tynan & DeMaster, 1997; Hovelsrud et al., 2008; Metcalf & Robards, 2008).

Habitat Modification Habitat modification in the form of less ice, thinner ice, and changes in ice characteristics present significant timing challenges to polar pinnipeds that rely on ice for resting, breeding, feeding, and refuge. Walrus depend on ice as a resting platform as they feed on the benthos of shallow continental shelves. As the ice recedes north of the continental shelf into the deeper Arctic Ocean, walrus are unable to dive to the depths where their benthic prey is located and return to resting locations (Metcalf & Robards, 2008; Moore & Huntington, 2008). This has caused more walrus to haul out on land closer to feeding areas, which puts them at greater risk for the spread of disease, disturbance, contamination, and being trampled in the crowded haul out regions (Kelly, 2001; Metcalf & Robards, 2008). Indirect impacts of habitat modifications could radically affect the acoustic ecology of polar pinnipeds. Van Parijs et al. (2004) indicated that interannual differences in ice conditions affected male mating strategies of bearded seals and the number of acoustic displays, hence influencing their behavioral ecology. Furthermore, uncharacteristically high concentrations of animals on land and decreasing ice resources add an additional selection pressure to the acoustic communication system of polar pinnipeds. Increases in noise level generated by the regional increase of animals in a haul out area will make it exceedingly more difficult for mothers and pups to establish and maintain acoustic contact. Additionally, the acoustic soundscape will change

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as ice conditions change (Figure 5). Periods of relative quiet during the Arctic winter when pack ice is present may become increasing rare (Figures 5 & 6). This quiet period overlaps with the time when many pinniped species are engaged in acoustic activity associated with mating (Van Parijs et al., 2001; Van Parijs, 2003; Van Opzeeland et al., 2010). How this will impact acoustic communication related to mating is unknown.

Ecosystem Alteration Ecosystem or regime shifts indicative of ecosystem alteration have already been observed in the sub-Arctic Bering Sea and could be a forecast of the future if ice continues to decrease in the Arctic. A time series of temperature measurements from the middle Bering Sea ice shelf indicates a warming trend, rather than interannual variability. This suggests that if climate continues to warm, the phytoplankton and zooplankton communities may shift from large to small taxa, which could strongly impact apex predators and the economies they support (Coyle et al., 2008). A regime shift observed in the Northern Bering Sea indicated a transition from a benthic driven ecosystem more characteristic of polar environments to a pelagic system more representative of temperate ecosystems (Grebmeier et al., 2006). The impacts of changing ice-plankton dynamics directly affect food availability for polar pinnipeds, but it also introduces competition from temperate marine mammal species. Seasonal and perennial ice has historically served as a physical barrier separating polar and temperate species (Stirling, 1997). Reduction in ice extent and thickness will create opportunities for more temperate species to migrate poleward and compete directly with polar species for resources while also introducing acoustic competition.

Stresses to Body Condition and Health Sea ice is a critical habitat that is closely linked to the health, body condition, and reproductive success of ice-obligate marine mammal species. Possible direct health effects of climate change for polar marine mammals include increased drownings and strandings as a result of sea ice platform loss (Burek et al., 2008). Indirect effects of climate change on marine mammal body condition and health may include increased pathogen

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transmission rates as a result of increasing haul out densities due to fewer suitable haul-outs (Lavigne & Schmitz, 1990). Given that vocalization numbers and patterns serve in many species as an indicator of fitness (Van Parijs, 2003), stress to body condition and health as an indirect or direct consequence of climate change may also affect acoustic behavior and eventually individual or population reproductive success. Walrus depend on the ice as a resting platform between foraging bouts. When the ice in the Arctic retreats further north, off the continental shelf and into deeper waters, walrus must travel further to feeding areas or shift to land for rest. During years with less ice and early ice retreats, the physical condition of Pacific walrus was reported to be poor (Pungowiyi, 2000; Metcalf & Robards, 2008). For ringed seals, early ice break-up and overall reductions in ice jeopardize reproductive success and pup health. During poor ice conditions, pups separated from their mothers were forced to enter the water prematurely, resulting in shorter nursing periods which led to slower growth rates and higher pup mortality (Smith & Harwood, 2001). In the Antarctic, similar projections have been made for Weddell seals that also rely on the pack ice for resting between feedings (Tynan et al., 2010). The increase in seal density and transit times between haul-out sites and feeding areas during periods with reduced ice could result in greater energy expenditure and competition for food leading to poor body condition. For the few Weddell seal colonies that are known to breed on sandy and rocky beaches (South Georgia, South Orkney Islands, e.g., Bonner, 1958; Vaughan, 1968), the availability of suitable ice for breeding might be less critical. Nevertheless, it has been suggested that early ice break-up of the ice pack around Signy Island had a negative effect on breeding (Vaughan, 1968). Too little is known on the behavior of these animals to draw further conclusions on the mechanisms behind this observation.

Increased Human Interactions The ice barrier has also played a critical role in separating polar species from human activities. Longer periods of open water availability allow for an increase in human use of the region for shipping/transportation, commercial fishing, recreation/tourism, and oil and gas energy exploitation. Increases in these activities introduce the potential for negative impact by increasing risks for entanglement, collision with vessels, by-catch, and pollution. Noise generated from anthropogenic activities may affect polar species through

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displacement from critical habitats (e.g., gray whales, Eschrichtius robustus: Stafford et al., 2007; Adelie penguins, Pygoscelis adeliae: Tenaza, 1971; beluga, Delphinapterus leucas: Finley et al., 1990), increased swim speed, longer dive durations, decreased inter-animal distance, and increased breathing synchrony (Richardson et al., 1995; Hastie et al., 2003). Additional acoustic impacts from anthropogenic noise increase the potential for masking of communication signals or other signals of biological importance (e.g., Richardson et al., 1995; Tyack, 2008). The threats cascading from global climate change in polar regions are complex and numerous. It is likely that impacts will first be observed in sub polar regions because dependence on seasonal ice presence plays less of a role in ecosystem dynamics. It is also likely that ecosystem impacts will occur sooner in the Arctic than the Antarctic. For these reasons, environmental and acoustical comparisons between polar and sub polar regions may provide insight into the future regarding ecosystem impacts on polar pinniped acoustic ecology.

CASE STUDY: ARCTIC PINNIPEDS Acoustic Data: Bering Sea Recordings for this case study were obtained with a Passive Aquatic Listener (PAL) deployed on a sub-surface NOAA FOCI mooring (M5) along the 70-m isobath of the eastern Bering Sea Shelf (Figure 7). Two years of data (2007-2009) from the central region of the Bering Sea were used to illustrate the seasonality of pagophilic Arctic pinnipeds. The PAL is an adaptive sub-sampling instrument that recorded at 2-5 minute intervals to provide consistent acoustic coverage over the 2 year period. The default sampling strategy was to record a 4.5 sec interval at 100 kHz every 5 minutes. This corresponded to a 1.5% duty cycle. When PAL real-time processing algorithms detected a period of increased acoustic activity or a signal of interest, the interval between samples was decreased to 2 minutes, resulting in a duty cycle as high as 4% during periods of high acoustic activity. Details related to the PAL sub-sampling strategy, detection thresholds, and probabilities of detection for marine mammals in the Bering Sea are found in Miksis-Olds et al. (2010). To examine pinniped acoustic behavior in relation to local ice cover, ice characteristics were obtained from the NWS Alaska Sea Ice Program.

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Percentage of ice cover and ice thickness above the mooring was estimated over a 20 x 20 km area.

Figure 7. Map of NOAA FOCI mooring locations along the 70-m isobaths in the Bering Sea. Variability in maximum sea ice extent was variable among years. Image compliments of P. Stabeno (NOAA PMEL).

Acoustic Activity of Arctic Pinnipeds in Relation to Ice Conditions Of the Arctic species (ringed, bearded, harp, hooded, grey, ribbon, Caspian, Baikal, spotted or Largha, harbor seal, and walrus), information on underwater vocalizations is known for 9 of the 11 species (see Stirling & Thomas, 2003 for a review). Currently there is no data on the aquatic vocalizations of Caspian and Baikal seals. The number of known underwater call types for Arctic pinniped species range from three in the ribbon seal

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(Miksis-Olds & Parks, in press) to 26 identified call types for the harp seal (Serrano, 2001). The majority of species, however, use only 4-6 call types (summarized in Van Parijs, 2003). Underwater vocalizations have been related to mating behavior in harp, spotted, harbor, bearded seals, and walrus. The latter three species exhibit stereotyped acoustic displays combined with visual and dive displays (Sjare & Stirling, 1996; Van Parijs et al., 2000; 2001). Of the five species where underwater vocalizations have been linked to mating behavior, information about the species‘ mating strategies is known for only three species. Male harbor seals perform acoustic displays as part of an aquatic lek system (Van Parijs et al., 1997; Hayes et al., 2004), whereas male bearded seals either defend aquatic territories or roam over a larger area (Van Parijs et al., 2004). Walrus are unique in that the males perform acoustic and dive displays as part of a polygynous mating system in which males defend females (female defense polygyny, Sjare & Stirling, 1996).

Figure 8. A spectrogram of a 4.5-sec interval recorded by a PAL at mooring site M5 in the Bering Sea on 26 April 2008. The spectrogram was produced from a recording sampled at 100 kHz using a 1024 point FFT with no overlap. Portions of the vocalizations of three different species overlap in time and frequency.

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In this case study, walrus, bearded, and ribbon seals were detected during the winter and spring of 2008 and 2009 in the Bering Sea (Figure 8). Further discussion in this case study will focus on these three species. Walrus and bearded seals have a circumpolar Arctic distribution, whereas ribbon seal are distributed between the northern boundary of the Chukchi and Beaufort Seas and southern boundary of the Sea of Japan. Walrus have been reported to vocalize underwater year round, but stereotyped displays have only been documented during the winter/spring breeding season. Bearded seal vocalizations were not detected in year round recordings from the Bering Sea and Kongsfjorden, Svalbard outside the breeding season from January-May in the Bering Sea and April-July in Kongsfjorden (Figure 9; Van Parijs et al., 2001; Miksis-Olds et al., 2010). Similarly, ribbon seal underwater vocalizations were not detected outside the winter/spring breeding season in the Bering Sea (Miksis-Olds & Parks, in press). Historically in the Bering Sea, ice can form as early as November and remain into late June (BEST, 2004). Seasonality of ice presence constrains pinniped breeding activity geographically and temporally for individuals mating in the sub polar waters of the Bering Sea compared to those mating in higher Arctic latitudes where access to open water is a driving factor constraining distribution and mating activities. Understanding the constraints imposed by ice, it is not surprising to detect multiple marine mammal species in the same general region vocalizing at the same time as they take advantage of the seasonal ice (Figures 8 & 9). Bearded, ribbon seals, bowhead whales, and walrus are the main contributors of biologic sound on the central Bering Sea Shelf from January to June (Figure 9; Miksis-Olds et al., 2010). Bowhead whales are included in the discussion of polar pinniped acoustics in this case study because they contribute significantly to the acoustic soundscape (Figure 9) and their sounds overlap in frequency with all three pinniped species present (Richardson et al., 1995).

Figure 9. Presence of bowhead whales and Arctic pinniped sounds in the Bering Sea from 2008-2009. Solid and dotted lines indicate percent ice cover and ice thickness (in inches), respectively. Acoustic presence of species does not correspond to a numerical value on the y axes. The species-specific symbols reflect daily acoustic presence and are separated in vertical space by species for easy visualization.

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The periods during which the various species are acoustically active in relation to the ice presence and characteristics suggests that ice dynamics may play different roles in different species. Bowhead whales are benthic feeders that migrate into the Bering Sea during winter ahead of the seasonal ice and follow the ice retreat back into Arctic waters in the late spring to capitalize on the high productivity associated with the ice edge (McRoy & Gowing 1974; McRoy & Goering, 1976; Hazard & Lowry, 1984; Moore & Reeves, 1993). Bowhead whale sounds were first detected prior to ice arrival and ended in late April/early May as the ice started to decline (Figure 9). In contrast, the first walrus and bearded seal vocalizations coincided with the seasonal arrival of ice. Both of these species are ice-obligate species that depend on ice for feeding, breeding, and resting. Walrus detections ended as the ice began to thin and retreat, whereas bearded seal vocalizations were detected until the ice completely retreated from the area. Onset of ribbon seal vocalizations was the latest and did not occur until the ice reached an approximate 20 inch thickness with 80% ice cover. These animals are considered an ice-associated species that rely on ice for parturition and lactation (Fedoseev, 2002). Ribbon seal and bearded seal detections ceased at the same time when the ice retreated from the area. Of particular note is the absence of bearded and ribbon seal vocalizations during a temporary ice retreat in March 2009 (Figure 9; Miksis-Olds et al., in prep). Without concurrent visual data, it was not possible to know whether these animals left the area in conjunction with the ice or whether vocal activity ceased while the animals remained in the area, but engaged in behaviors other than mating displays. Patterns of acoustic detection indicate that bearded seals and walrus are able to use ice with a wide range of characteristics, whereas ribbon seals require thicker ice with a greater percentage of ice cover. The peak periods of acoustic activity related to mating appear to be staggered in walrus and ribbon seals. The late peak detection of vocal activity by ribbon seals suggests that birthing and mating occur later in the season compared to walrus. Vocalizations associated with the breeding activity of bearded seals are evident throughout the season.

Inter-Specific Vocal Comparison of Arctic Pinnipeds Despite the temporal staggering of peak vocal activity by the different marine mammals present in the Bering Sea during winter/spring, there is a

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three month period when bowheads, bearded and ribbon seals, and walrus acoustically overlap (Figure 9). The potential for masking among the four soniferous marine mammal species is compounded by the sounds of ice as it advances, retreats, forms and melts. Biological sound production during the pinniped breeding season is louder and more variable in this region than at any other time of the year. There is little, if any, anthropogenic sound contributing to the overall soundscape during the ice-covered breeding season (Miksis-Olds et al., 2010) so the current acoustic environment is most likely representative of the environment in which the acoustic systems of polar pinnipeds evolved. In a noisy environment, some animals are able to compensate for noise to maintain an effective range of communication by increasing the amplitude, rate and duration of their sounds, modifying the frequency or structure of calls, or shifting the timing of vocal activity to quieter periods (Brumm & Slabbekoorn, 2005). Noise compensation behaviors are most useful when the noise is transient or short-term. One of the simplest noise compensation methods is to delay signaling until sound levels decrease (Tyack, 2008). In the case of ice breeding pinnipeds during the mating season when ice noise can be extremely loud, waiting for quiet periods to produce acoustic signals may not be advantageous. The fact that some species have been observed to call almost continuously during the mating season (e.g., Van Opzeeland et al., 2010) may reflect a strategy to increase signal detection probability under acoustic conditions in which quieter periods occur randomly. When noise (biotic, abiotic, or anthropogenic) is consistently and/or continuously present in the environment over long periods of time, selective pressure for vocal adaptation can lead to the development of new call structure features that are optimized for propagation given the noise and physical conditions of the environment (Morton, 1975; Tyack, 2008). For example, several bird species are known to exploit optimum frequencies to maximize long distance acoustic communication depending on the propagation characteristics of the environment (Marler & Slabbekoorn, 2004). Sound windows, or optimal frequencies for sound propagation, are different in forest, marsh, and grassland habitats. Birds have evolved acoustic signals with frequencies to maximize sound propagation in each habitat (Morton, 1975, Marten & Marler, 1977). A second selective pressure shaping signal structure and behavior is acoustic competition or overlap with nearby conspecifics or heterospecifics, also referred to as acoustic jamming (Tyack, 2008). This is a strong selective pressure influencing vocal activity in polar pinnipeds because their reliance

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on ice during the mating period results in multiple species performing acoustic displays at the same time in the same location. Harp seals have evolved jamming avoidance responses and can produce calls that differ in frequency by more than one-third octave to avoid acoustic overlap with nearby conspecifics (Serrano & Terhune, 2002). Bearded seals also have evolved similar anti-jamming behaviors and can alter the frequency of their vocalizations to avoid masking (Terhune, 1999). A comparison of the major call types and structures used by walrus, bearded seals and ribbon seals during the mating season in the Bering Sea suggested that acoustic competition among species during this time led to the evolution of signal types and vocal behaviors that minimize masking and jamming. Male bearded seals produce long tonal trills and sweeps with narrow frequency modulations during the mating season (Van Parijs et al., 2001), whereas the dominant call type of walrus during underwater acoustic displays is an impulsive, broadband series of knocks (Stirling et al., 1983; 1987). It is not known if bearded seal males are capable of shifting trill frequency to reduce jamming, but the frequency and time characteristics between the male bearded seal trills and the walrus knocks are different enough to minimize interference despite an overlap in the frequency range of calls by the two species (Figure 8). Ribbon seals produce intense, 1-4 sec downsweeps with harmonics during the breeding season that overlap in frequency with bearded seal trills, yet the bearded seal‘s long duration trills are clearly distinguishable from ribbon seal downsweeps (Figure 8). Similarly, long duration bearded seals trills bridge gaps between the units of bowhead song and thereby reduces masking and increases the probability of detection. Inter-specific differences in pinniped call structure and vocal behavior observed during the breeding season in the Bering Sea support the concept that biotic and abiotic factors shape vocal behavior resulting in the development of acoustic niches for species overlapping in time and space (Van Opzeeland et al., 2010). Pinniped vocal activity in the sub-Arctic waters of the Bering Sea during the ice-covered winter months shows similar acoustic adaptations to vocal patterns observed in the Arctic (Van Parijs, 2003; Van Opzeeland et al., 2008, 2010). The tight coupling of seasonal ice and vocal activity of polar pinnipeds in the Bering Sea also provides an appropriate comparison to Antarctic pinnipeds that function in an environment where a majority of the ice formed annually in winter melts during the austral summer.

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b) a)

4500 m 2500 m 2000 m 1000 m

c) Neumayer

Neumayer

Dronning Dronning Maud MaudLand Land

South South Pole Pole

Figure 10. a) Bathymetry map showing the location of PALAOA on the Eckström Ice Shelf (white star) and the location of the German Antarctic Neumayer Station II (black star). Inset image: map of Antarctica showing the location of Neumayer II, the German Antarctic Station (black star). b) Aerial picture of PALAOA on the ice shelf taken from the East (within Atka Bay), and c) the PALAOA station on the ice shelf. Photographs: AWI Logistics Department (b) and L. Kindermann (c).

CASE STUDY: ANTARCTIC PINNIPEDS Acoustic Data: PALAOA For this case study, underwater acoustic data collected by the PerenniAL Acoustic Observatory in the Antarctic Ocean (PALAOA) were used. The observatory is located on the Eckström Ice Shelf, located on the eastern Weddell Sea coast (70°31‘S, 8°13‘W, Figure 10, Boebel et al., 2006; Kindermann et al., 2008; Klinck, 2008). The ice shelf at this location has a thickness of about 100 m, with approximately 160 m of water between the base of the ice shelf and the ocean bottom. Continuous acoustic recordings are made year-round with two hydrophones suspended 80 m below the ice shelf. The hydrophones were lowered on their cables through two boreholes (separated by 300 m) and connected to the energetically autonomous PALAOA station. Recordings were digitized at 48 kHz/16 bit and encoded to a 192 kBit MP3 stream by a BARIX Instreamer device. The effective bandwidth of the recordings was 10 Hz to 15 kHz and the dynamic range was 60 dB to 150 dB re 1 µPa (see Kindermann et al., 2008 for more detailed information). The observatory has been operational since December 2005.

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Herein, acoustic data from PALAOA from 2006 were sub-sampled and manually analysed (see Van Opzeeland et al., 2010 for details on the subsampling and analysis procedure). To compare Antarctic pinniped acoustic presence with local ice cover, percentage of open water was calculated using ENVISAT ice cover data with a 6.25 x 6.25 km resolution (based on the algorithms described in Spreen et al., 2008). Determination of the monitoring radius around PALAOA was calculated was based on estimations of the approximate distance over which mid-frequency pinniped calls propagate (Van Opzeeland, 2010).

Acoustic Activity of Antarctic Pinnipeds in Relation to Ice Conditions Recordings from PALAOA contained vocalizations of all four Antarctic pinniped species (Weddell, leopard, Ross and crabeater seals) as well as sounds produced by various baleen whale species (Antarctic blue, Balaenoptera musculus intermedia, humpback and fin whales), odontocete species (e.g., killer and sperm whales, Physeter macrocephalus), ice sounds, and occasional vessel noise. The soundscape varied dramatically between seasons, with most biotic acoustic activity occurring in austral summer between October and February. Apart from ice sounds (e.g., calving of glaciers), the four pinniped species are the most prevalent sound sources in the soundscape during this period, with the periods of acoustic activity largely sequenced between species. Van Opzeeland et al. (2010) discussed how the four pinniped species might reduce acoustic interference by interspecific acoustic niche partitioning, suggesting that each species fills its own acoustic ecological niche. This case study discusses how the vocal behavior of two (Weddell and Ross seals) of the four pinniped species related to local ice conditions (Figure 11). Although Weddell seal calling activity was low in January and March, they were acoustically present in the PALAOA recordings throughout the year, except in February. Ross seals on the other hand, are only seasonally detected in the PALAOA recordings from December until February. Data on local ice conditions in a 40 km radius around PALAOA show that Weddell seal acoustic presence is strongly linked to the presence of fast ice, whereas Ross seal calls occur after the major ice break-up in December and are linked to the presence of pack ice. Furthermore, Ross seal calls occurred exclusively during the mating period, whereas Weddell seals call during the mating period and beyond.

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Call activity [arb. units] (%) coverage IceIce (%) coverage (%) Ice coverage Call activity [arb. units]

100

Pack-ice

80 60

Fast-ice

40 20 0

Jan 2006

Apr

Jul

Oct

Jan 2007

Mating

Figure 11. Acoustic presence of Weddell (blue) and Ross (green) seals in relation to ice conditions (40 km radius around PALAOA). Dashed bars in the bottom of the figure indicate the timing of mating for each species.

The two species have a very different behavioral ecology, with Weddell seals being a stationary species, present in the area almost the whole year. Weddell seals are thought to be a territorial species that uses its calls to defend underwater territories and attract mates (Bartsch et al., 1992; Stirling & Thomas, 2003; Rouget et al., 2007). The loud (source level up to 193 dB re 1 μPa), large (up to 34 underwater sound types) and complex Weddell seal vocal repertoire (Thomas & Kuechle, 1982) suggests that calls might be used in inter- and intrasexual communication with nearby breeding colonies and therefore are not constrained by signal propagation needs (Figure 12). The fact that Weddell seals are vocally active outside the period that mating takes place, might reflect that calls are produced by males that occupy underwater territories year-round (Rouget et al., 2007; Van Opzeeland et al., 2010). Males that remain in their territories might have an advantage over nonterritorial males or males that move away in winter, in that they are already resident when females arrive at the breeding area (Harcourt et al., 2007; 2008) or when the land-fast ice first forms (Rouget et al., 2007). The ability of males to defend underwater territories is however likely to be strongly dependent on the availability of a stable fast ice environment, providing spatially relatively fixed environmental features used to mark the territory and/or its boundaries (Miller, 2009). In Weddell seals, males are thought to maintain underwater territories around breathing holes in the fast ice, thereby optimizing exposure of females to their underwater vocal displays, which can easily be heard by females hauled out in the colony with their pups and as females pass through the male‘s underwater territory on their foraging dives (Hindell et al., 2002, Sato et al., 2002). Seasonal ice cover in Atka Bay where most of the Weddell seal colonies that PALAOA recorded are located – is usually lowest in February. The absence of Weddell seal vocalizations in the PALAOA recordings throughout February might therefore reflect a short

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period during which seals either cease to be vocally active or have moved because of fast-ice breakup (either actively or drifting passively on ice floes) and thereby were out of the recording range of the observatory. Preliminary analysis of data on satellite-tagged Weddell seals in the Atka Bay area in February seems to support the latter hypothesis (pers comm. H. Bornemann & J. Plötz). After the seasonal fast ice break-up in January-February, the ice sheet in Atka Bay builds up again in March and April. Preliminary analysis of multi-year recordings from PALAOA showed that the number of Weddell seal calls was substantially lower in a year which had a sudden drop in ice cover in April (Van Opzeeland, 2010). Possibly this drop in ice cover represented a late ice break-up within Atka Bay, which might have affected settlement of Weddell seal colonies, leading to fewer animals present and breeding (and thereby being vocally active) that year or to settlement in different areas.

Figure 12. Spectrogram of Weddell seal underwater vocalizations recorded by the PALAOA observatory. Recording from 14 December 2006. Spectrogram created using a 2048 point FFT, with no overlap.

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Thomas et al. (1987) reported that Weddell seals dramatically reduced the number of underwater vocalizations in mid December, a time when leopard seal and killer whale predators moved closer to their colonies. They suggested that the sudden decrease in vocal behavior was an anti-predation adaptation. In the PALAOA recordings, the number of Weddell seal calls also decreases with increasing leopard seal calling activity around midDecember (Van Opzeeland et al., 2010). However, virtually no leopard seal calls are present in the PALAOA recordings during the period of Weddell seal acoustic absence in February. Furthermore, preliminary analysis suggests that killer whales are mainly (acoustically) present at PALAOA in January. The acoustic absence of Weddell seals in February in Atka Bay is therefore more likely to relate to changes in ice conditions than an antipredation strategy. In contrast to the stationary Weddell seal, Ross seals are a migratory species and are thought to be solitary and use their underwater calls only during the mating season to attract potential mates (Van Opzeeland et al., 2010). Their stereotyped calls and medium-sized call repertoire likely also reflect that their calls are used to communicate over larger distances (Figure 13, Rogers, 2003; Van Opzeeland et al., 2010). Preliminary analysis of multiyear data from PALAOA showed that the timing of Ross seal acoustic presence in the PALAOA recordings corresponds each year to the period when major ice break-up occurs in the area in austral summer and pack ice drifts in (reflected by the second peak in ice cover in January, Figure 11; Van Opzeeland, 2010). Years with lower ice cover in January resulted in less Ross seal call activity. The association between Ross seal acoustic activity and the presence and amount of pack ice in the area off PALAOA may either suggest that seals enter the area with the ice floes (either actively swimming or passively drifting on floes), or are already present in the area but await the availability of suitable ice before starting their vocal displays. A study involving six satellite-tagged Ross seals, showed that they migrate from coastal waters off the Antarctic continent to the pack ice edge near the Polar Front around February, where they stayed until October, then migrated south, reaching Antarctic coastal waters around mid December (Blix & Nordøy, 2007). This suggests that Ross seals are more likely to move into an area with the ice, rather than that they are already present in the area. Because so little is known on the biology of this species, much remains to be speculated with respect to Ross seal behavior. Virtually nothing is known on the specific characteristics of the pack ice type which Ross seals use for breeding. Siniff et al. (2008) suggested that floe size was likely to be of importance to this

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species when selecting their sea ice habitat for pupping and molt. Data from acoustic recorders deployed in offshore areas are currently under analysis to explore to what extent the coastal area off PALAOA represents rare or specific sea ice habitats where Ross seals aggregate to breed or if similar patterns can be observed in other offshore sea ice areas as well.

Figure 13. Spectrogram of Ross seal vocalizations recorded by the PALAOA observatory. Recording from 25 January 2007. Spectrogram created using a 2048 point FFT, with no overlap.

CONCLUSION What Can We Learn by Taking a Bi-Polar/Comparative Approach? The two case studies presented illustrate that certain pinniped species restrict their acoustic activity to periods with suitable ice conditions. While bearded seals and walrus seem able to use ice with a wider range of

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characteristics, ribbon, Weddell and Ross seals are only acoustically present when specific ice types dominate the region in which recordings were made. These species may therefore be particularly vulnerable to changes in their sea ice habitat and may serve as ‗sentinel species‘ to remotely monitor ice conditions. Nevertheless, it is important to keep in mind that some species only produce sound during specific times of the year (i.e., the mating season), limiting the period during which the presence of the these species in certain areas can be acoustically monitored. Furthermore, variation in other habitat factors, such as the availability of suitable feeding habitat near sea ice haulouts, is likely to cause strong local differences in marine mammal vocal responses to changes in sea ice conditions Acoustic techniques provide an important remote sensing tool with which the response to changes in ice habitats can be monitored. Given that logistical constraints limit access to polar regions for certain periods of the year, acoustic tools are of particular value to monitor species in these remote regions. It is important to note, however, that in some cases additional observation methods are necessary to assess if animals are present, but not vocalizing or have left the area, such as in the case of the void in bearded, ribbon and Weddell seal vocalizations during a period of reduced ice cover, as was discussed in both case studies. Such methods may include tagging animals with acoustic and/or GPS tags and visual surveys, although the use of these methods in polar regions is often constrained to some extent by environmental conditions (e.g., Van Opzeeland, et al., 2008) and the molting cycle of adults (i.e., satellite tags attached during the breeding season when animals are easiest to catch will fall off during the molt a few months later). Furthermore, it can be concluded that ice conditions are an important driver within the acoustic ecology of pinnipeds in polar regions. In both Arctic and Antarctic pinnipeds, the constraints imposed by the linkages between ice and breeding cycles are likely to force multiple species to time their acoustic activity within the same period. The Arctic case study provides an example of how the acoustic characteristics of the calls of the various marine mammals have likely evolved to prevent masking and jamming of signals. In the Antarctic, the broad frequency bandwidth of Ross seal calls and the relatively high amplitude of leopard seal calls seems to temporarily ‗block‘ the acoustic space for other species during the distinct period when each of these species is vocally active. This might be another mechanism that has evolved to prevent inter-specific acoustic interference (Van Opzeeland et al., 2010). Complete separation of acoustic activity between species might on the other hand not be possible because of biotic (e.g., prey availiability)

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and/or abiotic factors (e.g., required communication range and transmission properties of acoustic environment restrict frequency shifting of vocalizations). Acoustic niche partitioning has previously been observed in various other animal taxa (e.g., Ficken & Hailma, 1974; Gerhardt, 1994), and it is likely that this phenomenon is also of importance to acoustic communication in marine mammals. In light of climate change, changes in ice conditions may indirectly lead to changes in the timing of mating, and hence acoustic activity. If acoustic niche partitioning is one of the drivers behind the species-specific timing of acoustic activity, changes in ice conditions may affect the acoustic ecology of polar pinnipeds through changes in the level of inter-specific acoustic interference. An additional effect of changes in ice conditions on polar pinniped acoustic ecology, may involve changes in the presence and acoustic characteristics of ice sounds (i.e., abiotic soundscape) during the periods when polar pinnipeds are vocally active. Both changes in the level of inter-specific interference and changes in the abiotic soundscape are likely to affect species differently, depending on the typical communication range and acoustic characteristics of the species. Furthermore, the fact that some polar pinniped species exhibit geographical variation in their repertoires (e.g., Weddell, leopard and bearded seals, Thomas & Stirling, 1983; Thomas & Golladay, 1995; Risch et al., 2008) suggests that these species are vocally flexible and innovative in forming new or modified sounds. Some polar pinniped species might therefore be capable of adapting their calls in response to long-term changes in their acoustic environment. Before further inferences on the potential effects of changes in ice conditions on the acoustic ecology of polar pinnipeds can be made, longterm acoustic data are necessary to develop a baseline understanding of the key processes and factors that shape species-specific vocal behavior. By taking a comparative approach between pinnipeds inhabiting Arctic and Antarctic regions, it becomes possible to better understand how behavioral ecological characteristics that evolved independently are linked to specific ice conditions. Such knowledge can contribute to a better understanding of how pinnipeds in polar regions use their habitat and how potential changes may affect them. Monitoring ecosystem response from the perspective of acoustic ecology in the Arctic, may provide insight into expected changes in the Antarctic, as Arctic conditions are projected to change sooner and more quickly than Antarctic conditions. Similarly, the investigation of changes in sub polar regions may provide important information for predictive capabilities on how ambient sound may change in

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the high latitudes of the polar regions and hence impact pinniped acoustic ecology.

Outlook In terms of acoustic ecology, the increase in ambient sound due to increased human activity in polar areas, along with changes in the abiotic and biotic soundscape (resulting from changes in ice sounds, timing of vocal activity, and/or increased vocal presence of migrant species), are likely to be of most concern. Adaptive responses are generally slow and can only accommodate gradual environmental changes, leaving the ability of polar pinnipeds to cope with changes in their environment dependent on the time scale over which changes occur. Coupled with the reduced availability of suitable ice platforms for resting, molt and breeding, changes in polar pinniped acoustic ecology could lead to conditions of poor body condition, and eventually reduced reproductive success with consequences for species at the population level. It is not unrealistic to expect ecosystem effects to impact acoustic ecology, as the transition of a benthic to pelagic dominated ecosystem associated with ice reduction and the loss of predictable polynyas will impact the distribution of benthic feeding pinnipeds and their vocal presence. Long term data sets on ambient sound and polar pinniped acoustic behavior are necessary to seek correlations with extreme ice and weather conditions in order to understand how climate change may affect these animals. Fortunately, recent advances and ongoing development of acoustic techniques have enabled collection of acoustic data over long time spans in the remotest of polar habitats. Furthermore, progress and development of techniques with which marine mammal density information can be extracted from acoustic data will be an important tool for abundance estimations in polar regions in the near future. As a final note, it is important for future studies that investigators realize that populations or in the case of many Antarctic pinnipeds even local breeding groups (Davis et al., 2008) are a more appropriate unit than species. In polar regions, the effects of climate change may vary dramatically between areas, causing responses of animals to vary with local environmental conditions.

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ACKNOWLEDGMENTS The PALAOA project was conceived and implemented by Olaf Boebel, Lars Kindermann and Holger Klinck with continuous support from the AWI logistics department, the Neumayer overwintering teams 2005-2010, FIELAX and Reederei Laeisz. The PALAOA project was partly funded by the Bremerhavener Gesellschaft für Innovationsförderung und Stadtentwicklung (BIS). We would like to thank Phyllis Stabeno, Carol Dewitt, Bill Floering, and Rick Miller from PMEL/NOAA for including our acoustics instruments on the NOAA FOCI moorings and for providing mooring logistical support. Thanks are also extended to the captain and crews of the NOAA ships and contract vessels that deployed and retrieved the mooring instruments. Funding for the Bering Sea project was provided by the ONR Marine Mammal Program under Award Numbers N000140810391 and N000140810394.

Reviewed by John M. Terhune (University of New Brunswick) and Jeanette A. Thomas (Western Illinois University-Quad Cities).

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In: Aquatic Animals Editor: David L. Eder

ISBN: 978-1-61470-123-1 © 2012 Nova Science Publishers, Inc.

Chapter 2

IMMUNOTOXICOLOGICAL REACTIVITY OF HEMOCYTE OF JUVENILE MUDCRAB OF SUNDARBANS BIOSPHERE RESERVE OF INDIA Sajal Ray*, Sanjib Saha and Mitali Ray Aquatic Toxicology Laboratory, Department of Zoology, University of Calcutta, 35 Ballygunge Circular Road, Kolkata – 700019, West Bengal, India

ABSTRACT Sundarbans Biosphere Reserve of India supports a wide range of flora and fauna. Ecologically, this reserve is a mangrove forest with numerous estuarine creeks. Mudcrab Scylla serrata (Crustacea: Decapoda) is an important member of estuarine habitat of Sundarbans. Indiscriminate collection of juvenile crab seeds and contamination of estuarine water pose a serious threat for this aquatic animal. Generally invertebrates including mudcrab lack adaptive immune system and depend on innate immune defense to combat pathogen and toxin . Cellular immune response involved the recruitment of immunocompetent hemocytes capable of performing adhesion, aggregation, phagocytosis, generation of intracellular cytotoxic molecules etc. In this present study, spectrum of innate immunological * Corresponding author. E- mail: [email protected]

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Sajal Ray, Sanjib Saha and Mitali Ray reactivity was examined in juvenile mudcrab under the exposure of arsenic, a toxic metalloid of aquatic environment. Experimental juveniles were exposed to 1, 2 and 3 ppm of sodium arsenite for 1, 2, 3 and 4 days in controlled laboratory condition. Total hemocyte count is a stress indicator of pollution which was enumerated both in the control and arsenic exposed S. serrata. Crabs exposed to sodium arsenite expressed an alteration in total hemocyte density. Aggregation and adhesion of hemocytes are metabolic behaviours whereas phagocytosis is a classical immunological response of hemocytes and arsenic exposure yields a dose responsive decrease in aggregation, adhesion and phagocytosis against the control indicating perturbation of blood cell hemeostasis, alteration of hemocyte surface characteristics and immune suppression. Capability of generation of cytotoxic molecules like superoxide anion and nitric oxide are reported as a potential strategy of immunological defense against pathogen invasion. Disruption of hemocyte density, adhesion, aggregation, phagocytosis and increment of cytotoxic response may lead to a state of alteration of immunity. Arsenic induced alteration of immune status may impart a state of vulnerability in juvenile crab inhabiting the toxic aquatic environment. Such a situation may reduce the ecological fitness of mudcrab to combat against the toxic challenge caused by arsenic pollution.

Keywords: Arsenic, juvenile, hemocyte, mudcrab

INTRODUCTION The Sundarbans delta is the world‘s largest mangrove forest belt covering 9360 square km. of total area of India and Bangladesh. The Sundarbans is the first mangrove forest included in the status of ‗World Heritage‘ in 1987 by IUCN. Mudflats of Sundarbans estuary are characteristic of low tide and high tide. This mudflat are exposed during low tide and submerged during high tide and support a wide diversity of crabs of 67 species including Scylla serrata (Naskar and Ghosh, 1989; Chaudhuri and Choudhury, 1994; Ali et al., 2004). Sundarbans mangrove ecosystem is a dynamic, productive and diversified one that is currently facing several types of stresses of anthropogenic and natural origin. Deforestation, over exploitation of natural resources, soil erosion, deposition of heavy metals, pesticides, extension of population, aquacultural practices, discharge of untreated sewage and oil spillage from motorized vessels are creating ecological crises and stress on Sundarbans Biosphere Reserve (Sarkar et al.,

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2002 and 2004; Ghosh, 2007). The main impact and crisis of mudcrab farming in India resulted indiscriminate wild seed collection, absence of scientific practice of aquaculture and change of climate. As a result, hatchery production for S. serrata appears to be unreliable. Excessive harvesting of wild seed may deplete wild crab population of this area. Natural habitat of S. serrata bears the environmental threat due to exposure of diverse hazardous xenobiotics, including arsenic from different natural and anthropological sources (Krishnaja et al., 1987; Gomez- Caminero, 2001). Hemocytes are the chief immuno effector blood cells of the invertebrates that perform diverse immunological as well as physiological functions including aggregation, adhesion, and phagocytosis under the exposure of diverse physiological challenges (Bodammer, 1978; Saha et al., 2007; 2008; 2009 and 2010). Circulating hemocyte density is considered as a valuable tool for biomarking the physiological stress of invertebrates exposed to xenobiotics including arsenic (Saha and Ray, 2006; Chakraborty et al., 2008; Saha et al., 2007 and 2009). Aggregation is an important cellular behaviour involved in recognition of self as well as nonself foreign surfaces, clot formation at wound site, encapsulation reaction and intercellular communication (Kenney et al., 1972; Takahashi et at., 1995; Chen and Bayne, 1995). Recognition and adhesion of hemocytes to nonself source is an important immune response against invading pathogens. Attachment, spreading and migration of hemocytes are the physiological sequences during nonself surface adhesion process (Armstrong, 1980). During wound repairment process, cellular clot of hemocytes are formed and adhere to the surfaces of the wound (Johansson and Soderhall, 1988). Phagocytosis of nonself particulates is a classical innate immune response in invertebrates by which cells engulf relatively smaller particles, digest and release the remnants by the process of exocytosis from its body. This is an essential reaction for host defense against infectious microorganisms at the wound site and for the clearance of apoptotic cells generated during development (Ratcliffe, 1985). Phagocytosis is initiated by the recognition of target particle by a phagocytic cell characterised by binding, ingestion and clearance of pathogen through the blood cells that circulate widely in the body cavity and migrate by chemotactic behaviour towards the foreign invaders (Takahashi and Mori, 2000). To eliminate the phagocyted particulates from body, blood cells release cytotoxic chemical compounds as toxicological response (Anderson et al., 1992). Generation of cytotoxic molecules is reported as an innate immunological response. Superoxide anions are the first reactive oxygen intermediate that generate during this process and subsequent reactions produce hydrogen peroxide,

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hydroxyl radicals, singlet oxygen and finally water (Rodriguez and Moullac, 2000). Strategy of oxygen radical mediated killing is based on the premises of toxicity evolved by high concentration of molecular oxygen. Superoxide anions are extremely toxic, powerful and hyperactive agents and are capable of creating damage at cellular level. Intrahemocyte superoxide anion is generated at cell surface when the consumption of oxygen in cell is increased due to respiratory burst following the contact of the phagocyte to microorganisms (Munoz et al., 2000). Within mitochondrial membrane superoxide anion remains stable and can diffuse across the membrane in concentrationdependent manner but at extremely slow rate and superoxide dismutase reduces the steady-state concentration of superoxide anion by several orders of magnitude (Lesser, 2006). In invertebrates, mainly hyaline hemocytes are responsible for the production of superoxide anions and generation is spectrophotometrically detected following the principle of nitro blue tetrazolium reduction to blue formazan (Anderson et al., 1992; Munoz et al.2000). In Crustacea, pathogens are melanized by deposition of melanin around them. Melanization is usually observed by blackening of the host and pathogen in the hemolymph as black spots on the cuticle which often inhibits the growth of pathogens (Cerenius and Soderhall, 2004). The prophenoloxidase activating system comprises of an enzyme cascade leading to activation of prophenoloxidase and other compounds with related activities. The zymogen prophenoloxidase of is activated through the cleavage to phenoloxidase. Phenoloxidase is a copper containing protein that is activated by limited proteolysis and it catalyzes both the O- hydroxylation of monophenols to diphenols and oxidizes diphenols to quinines, which can polymerize non-enzymatically to insoluble melanin. Phenoloxidase is sticky in nature and can adhere to the surface of pathogens which, lead to melanize that can seal the wound, harden, darken the post-molt carapace and minimize infections (Nappi and Christensen, 2005; Tanner et al., 2006). This system can be activated and elicited by microbial cell wall components through recognition protein or receptor present in the hemolymph. Several pattern recognition proteins have been identified that are participating in immunorecognition and bind to microbial carbohydrate. Such complexes in the hemocyte surface induce different immune reactions such as spreading and degranulation of granulocytes. Degranulation process of hemocyte also occurs by other factors or compounds such as endogenous factors produced by damaged tissues (Soderhall and Cerenius, 1998). The activation of prophenoloxidase is regulated by serine protease cascades that cleave of the pro-form of the activating enzyme into active prophenoloxidase

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(Sritunyalucksana and Soderhall, 2000; Cerenius and Soderhall, 2004). To avoid the deleterious effects of active components of the prophenoloxidase system proteinase inhibitors play an important role to prevent the over activation, which can directly inhibit the activity of phenoloxidase (Aspan et al., 1990). Crustaceans innate immune system involves prophenoloxidase activation mechanism, exhibited by the granular hemocyte (Soderhall and Smith, 1983; Sung et al., 1998; Sahoo et al., 2005). Immunoreactive status of hemocyte, the chief immunoeffector cells of juvenile crabs was examined in depth under the experimental exposure of sodium arsenite. Total hemocyte count, aggregation, adhesion and phagocytosis of nonself particulates were studied under the exposure of sodium arsenite in controlled laboratory conditions in juvenile S. serrata.

MATERIALS AND METHODS 1. Collection, Transportation and Maintenance of S. Serrata Live intermolt adult specimens of S. serrata weighing 20-25 gm were collected throughout the year from selected habitats of south 24 Parganas of West Bengal, India which are relatively free from arsenic contamination. Length and breadth of the carapace of experimental animals were 2 ± 0.5 cm and 3 ± 0.5 cm respectively. The crabs were carried alive in jute bags and transported immediately to the Aquatic Toxicology Laboratory of University of Calcutta, West Bengal, India. In the laboratory live crabs were kept in large rectangular glass aquaria containing simulated marine salt water in controlled laboratory condition under constant aeration (Ali et al., 2004; Saha and Ray, 2006). Live animals were maintained in arsenic free distilled water in which marine salt was dissolved and fed with fresh flesh of prawn and molluscs once a day. The water of glass aquaria was replenished to avoid residual toxicity. Animals were acclimatized in laboratory conditions for 6-8 days prior to experimentation. Water quality parameters were kept constant. Temperature and acidity were measured using a pH meter, dissolve oxygen, hardness and salinity were measured by standard protocols (Heasman and Fielder, 1983). A photoperiod of 12 hr light-12 hr darkness was provided. High tide and low tide actions were simulated in laboratory static water condition by increasing and decreasing the water level in every 12 hours duration.

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2. Treatment with Sodium Arsenite After proper acclimation to the laboratory conditions, acute toxicity study was carried out to determine the lethal (LC100), median lethal (LC50) and safe sublethal (LC0) concentrations of sodium arsenite in S. serrata. Determination of LC50 was done by ‗Behrens-Karber method‘ after Klassen (1991) where as LC100 and LC0 were determined after Shibu Vardhannan and Radhakrishnan (2002) respectively. The crabs were divided into multiple groups and each group contained 10 acclimatized crabs of same age and sex in triplicate which were introduced in the glass aquaria containing 10 liters of water. From the stock solution desired concentrations of sodium arsenite were prepared for determining the toxicity of S. serrata. During the experiment dead crabs were discarded immediately from the treatment. For in vivo treatment, crabs were exposed to 1, 2 and 3 ppm of sodium arsenite (E. Merck, Germany; 99% pure; CAS number 7784-46-5) in double distilled static water by dissolving marine salt into borosilicate glass containers (Borosil, India). During treatment mouth of test concentration were partially covered by glass lids to prevent atmospheric desiccation. Height of the water column never exceeded a depth of two inches both for control and treatment states. Water was replenished at every 12 hours along with fresh concentration of sodium arsenite water and crabs of all the containers were fitted with electrically operated aerator to facilitate artificial supply of oxygen. During treatment the animals were routinely cheeked for mortality and morbidity. During laboratory maintenance control and treated crabs were fed with fresh small molluscans flesh (Bellamya bengalensis). Residual food part and dead individual were instantly removed from treatment jars to avoid residual toxicity.

3. Collection of Hemolymph Fresh hemolymph (~ 2 ml) was collected aseptically from the base of one of the second walking legs S. serrata using a sterile syringe fitted with a 23-gauge needle citrate EDTA buffer (0.45 M NaCl; 0.1 M glucose; 30 mM trisodium citrate; 20 mM citric acid; 100 mM EDTA, pH 4.6) was used as anticoagulant (Bell and Smith, 1994). Care was taken to avoid mixing of tissues during the puncture and collection of the hemolymph sample was done directly from the walking legs. After collection hemolymph was stored in prechilled sterile glass vials and examined under microscope (Axiostar

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Zeiss, Germany). Suspended hemocytes were sedimented by cold centrifugation (Hermle, Germany) of the hemolymph at 3000 rpm for 5 minutes and cell density was adjusted with uniform volume of sterile phosphate buffered saline (Saha et al., 2007 and 2009). Free and live hemocytes in suspension were examined and screened under inverted microscope (Axiovert, Zeiss, Germany).

4. Testing of Cell Viability Cell viability was tested by staining the cell with trypan blue following the principle of dye exclusion. For this experiment 100 µl of freshly collected hemolymph was mixed on a glass slide with 100 µl of a 1.2% solution of trypan blue modified after Hose et al. (1990). The ratio of blue stained and colourless cells were determined microscopically by estimating the percentage of these stained and unstained cells (Sauve et al., 2002). At least six hundred cells were screened for each specimen in this experiment. Experiments were carried out with cell suspension containing more than 95 % of viable hemocytes.

5. Determination of Total Hemocyte Count (THC) Total count of hemocyte was carried out by fresh hemolymph collected from live adult specimens. Total count is made by using improved Neubauer hemocytometer (Germany) under light microscope (Axiostar Zeiss, Germany) after Ravindranath (1977). Total count was expressed as number of 106 cells per 1 ml of hemolymph. Generated data was an average of at least five observations for each experiment (Saha and Ray, 2006; Saha et al., 2007 and 2009).

6. Aggregation Assay Fixed volumes of hemolymph with uniform density of hemocytes were placed chilled microfuge tubes in by using autopipette. One of the samples was fixed by 100 µl of 2.5% glutaraldehyde and then vortexed for 1minute. ‗No aggregation‘ value and degree of hemocyte aggregation were obtained after Chen and Bayne (1995).

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7. Non-Self Surface Adhesion Assay 100 µ of hemocyte suspension was placed on cover slip carefully and were allowed to settle over cover slips. Cell suspension was incubated in sterile humid chamber at 37oC for 150 minutes for completion of adherence of hemocytes in vitro. After 150 minutes of incubation, supernatant was removed from post incubated sample by micropipette and transferred into microfuge tubes. Adhered cells were subjected to gentle jetting of phosphate buffer saline to collect the nonadherent cells settled due to gravity. Population of adherent and nonadherent hemocytes were subsequently processed, fixed, stained and enumeration (Guria and Ray, 2003; Saha et al., 2008).

8. Assay of Phagocytic Response Under the Challenged Yeast For determination of phagocytic response under sodium arsenite in vivo, fixed numbers of untreated and treated hemocytes were challenged with freshly cultured yeast at an optimal phagocytic ratio of (1:10). Cell mixture was placed over cover slips and allowed to settle over cover slips. Cell suspension was incubated in sterile moist chamber at 37oC for 6 hours for completion of phagocytosis in vitro. Cells were subsequently processed, fixed and stained for microscopic observation and phagocytic index (PI) was determined (Saha et al., 2008).

9. Activity of Superoxide Anion (O2-) Generation of intrahemocyte superoxide anion was estimated spectrophotometrically (Shimadzu UV – 1700, Japan) following the principle of nitro blue tetrazolium reduction reaction modified after Song and Hsieh (1994). Hemocytes were collected were sedimented by cold centrifugation at 3000 rpm for 5 minutes. Hemocyte density was adjusted with sterile phosphate buffered saline. The cells were stained with nitro bluetetrazolium (NBT, SRL, India) solution (0.03%) and vortexed and incubated for 30 minutes at 37oC. The reaction was stopped by adding 1ml of absolute methanol. Proper washing with 70% methanol, the hemocytes were dried and coated with a solution of potassium hydroxide (Merck, India) and DMSO (Merck, India) to dissolve the cytoplasmic formazan. The dissolved

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cytoplasmic formazan was spectrophotometrically measured at 630 nm and activity of superoxide anion was expressed as OD / minute / 106 cells.

10. Activity of Nitric Oxide (NO) Generation of intrahemocyte nitric oxide was assayed spectrophotometrically (Shimadzu UV – 1700, Japan) at 550 nm following the principle of Griess reaction (Green et al., 1982) using sodium nitrite as standard to determine molecular molar concentration of nitrite in the sample. Reaction process involved optical measurement of pink coloured nitrite, stable and non-volatile breakdown product of nitric oxide. Hemocytes were collected and suspended hemocytes were sedimented by cold centrifugation at 3000 rpm for 5 minutes. Hemocyte density was adjusted with 1phosphate buffered solution. For preparation of Griess reagent 0.1% N-naphthylethylene-diamine (SRL, India) was dissolved in 2.5% phosphoric acid (Merck, India) and finally 1% sulphanilamide (SRL, India) solution was mixed with above N-naphthyl-ethylene-diamine – acid solution. Fixed number of sedimented hemocytes were treated with Griess reagent and incubated for 30 minutes at 370C. This reaction system is capable of generating nitrite in biological materials including cell suspension. The molar concentration of nitrite in the sample was determined by using sodium nitrite as standard. Activity of nitric oxide generation was expressed as µM nitrite / minute / 106 cells.

11. Activity of Phenoloxidase (PO) Estimation of activity of phenoloxidase enzyme was carried out spectrophotometrically (Shimadzu UV – 1700, Japan) by monitoring the rate of formation of dopachrome from L-3, 4 DOPA (Tanner et al., 2006). After collection of hemolymph, hemocytes were washed for three times and sedimented at cold and hemocyte density was adjusted with. Sedimented hemocytes were treated with Triton X-100 for 30 minutes at 40C and hemocytes were routinely checked for completion of cell lyses. Hemocyte lysate supernatant was collected carefully as centrifuged membrane debris. Estimation involved incubation of hemocyte lysate with 10% sodium dodecyl sulfate (Merck, India) solution for 15 minutes at 250C. Postincubated sample was reacted with L-3,4 dihydroxy phenyl alanine (SRL, India). For

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preparation of L-DOPA solution 2mg of DOPA was dissolved in 2 ml of sodium phosphate solution (pH 7.5, 100 mM). For monitoring the rate of formation of dopachrome, sample mixture in cuvette was instantly capped, inverted and screened for change in optical density at 475 nm for 5 minutes. Activity of phenoloxidase was recorded as an average of increment in OD/minute with at 5 minutes of interval. Enzyme activity was expressed as an increase in absorbance of 0.001 OD / min / mg protein.

12. Estimation of Protein Protein was estimated spectrophotometrically (Shimadzu UV-1700, Japan) following the method of Lowry et al., (1951) using bovine serum albumin (BSA) as the standard.

13. Statistical Analyses All experiments were repeated for 3 times and values are expressed as mean ± SD. The data are compared in paired Fisher t-test and symbols indicate values that are significantly different from controls. P < 0.05.

RESULT AND DISCUSSION Median lethal concentration for 96 hours, safe concentration and safe application rate of sodium arsenite in juvenile S. serrata were 10 ppm, 0.6 ppm, 6.6 ppm respectively in static water environment. Mean total hemocyte count (THC) was 25.9 ± 2.3 x 105 cells/ml and alterations of THC were studied against 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours of exposure in control laboratory condition (Figure 1). The highest hemocyte count was recorded against 3 ppm of sodium arsenite exposure for 96 hours as 134.23 x 105 cells/ml and the lowest hemocyte count was estimated against 1 ppm of sodium arsenite for 24 hours exposure as 42.32 x 105 cells/ml (Figure 1). Exposure of juvenile crab to 1, 2 and 3 ppm for 48 and 72 hours of span exhibited intermediate hemocyte counts against control value 25.98 x 105 cells/ml (Figure 1). Alteration in hemocyte aggregation was recorded in depth under the exposure of 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours against control (Figure 2 and 3). Kinetics of

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aggregation resulted a dose dependent increase of free cells under the exposure of arsenite. Highest and lowest values of aggregation of hemocyte were estimated under the exposure of 1 ppm of arsenite for 24 hours as 80.11 % and 3 ppm of arsenite for 96 hours as 19.35 % against control value 90.1 % (Figure 3). Kinetics of nonself surface adhesion indicated a steady decrease of adherence of hemocytes at an interval of 150 minutes under the exposure of arsenite against 1, 2 and 3 ppm for 24, 48, 72 and 96 hours (Figure 4 and 5). Hemocytes collected from S. serrata exposed to 2 ppm of arsenite for 24, 48, 72 and 96 hours expressed intermediate values of adherent cell percentage. Highest and lowest inhibitions were reported against 3 ppm of sodium arsenite for 96 hours and 1 ppm for 24 hours arsenite exposure respectively (Figure 5). The pattern of kinetics of hemocyte adherence was identical in all the experiment. Impairment of phagocytic response was recorded against all concentration of sodium arsenite tested. Phagocytic responses of hemocytes were studied under the challenge of cultured yeast particle in vitro (Figure 6). Exposure to arsenic yielded a dose dependent decrease in phagocytic index (PI) under the challenge of yeast upto 96 hours against the control (Figure 7). Highest and lowest inhibitions of phagocytic response were determined as phagocytic index against 3 ppm for 96 hours and 1 ppm for 24 hours respectively as determined (Figure 7). Generations of superoxide anion were increased in hemocytes of the specimen exposed to sodium arsenite and activity of superoxide anion was highest against 3 ppm of arsenite for 96 hours of exposure (Figure 8). Gradual increase of intrahemocyte superoxide anion were recorded against 1,2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours and expressed a dose dependent response (Figure 8). This occurred due to possible damage in the cytoarchitecture of hemocytes. Exposure of juvenile mudcrab to 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours resulted in a dose dependent pattern of increase of intrahemocyte nitric oxide (Figure 9). Highest activity of nitric oxide was recorded in animals exposed to 3 ppm of sodium arsenite for 96 hours (Figure 9). Exposure of juvenile S. serrata to 1, 2 and 3 ppm for 48 and 72 hours of span exhibited intermediate generation of nitric oxide against control value (Figure 9). Activity of phenoloxidase was estimated from 24 to 96 hours of exposure of sodium arsenite (Figure 10). The highest activity of phenoloxidase was recorded against 3 ppm of arsenite for 96 hours of exposure. Gradual increase in the activity of phenoloxidase was recorded against all treatment sets (Figure 10). Treatment by arsenite resulted a dose dependent increase in the activity of intrahemocyte phenoloxidase indicating toxicity.

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Figure 1. Modulation of total hemocyte count (THC) of juvenile S. serrata under exposure of 1, 2, 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours. Data expressed as mean ± SD. P < 0.05. (n = 3).

Figure 2. Aggregating tendency of hemocytes of juvenile S. serrata stained with H&E. 100x.

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Figure 3. Modulation of aggregation of hemocyte of juvenile S. serrata under exposure of 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours. Data expressed as mean percentage (n = 3).

Figure 4. Glass surface adherence of hemocytes of S. serrata stained with H&E, 400x.

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Figure 5. Alteration of non-self surface adhesion of hemocyte of juvenile S. serrata under exposure of 1, 2 and 3 ppm of sodium arsenite (in vitro) for 24, 48, 72 and 96 hours. Data expressed as percentage (n = 3).

Nucleus

Cytoplasmic extension Yeast

Figure 6. Phagocytic response of hemocytes challenge by yeast in short term cell culture and stained with H & E. 1000x.

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Figure 7. Dynamics of phagocytic response of hemocyte of juvenile S. serrata under challenged of yeast exposed to 1, 2 and 3 ppm of sodium arsenite (in vivo) for 24, 48, 72 and 96 hours. Data expressed as mean ± standard deviation. P < 0.05. (n= 3).

Figure 8. Alteration of intrahemocyte superoxide anion (O-2) generation of juvenile S. serrata under exposure of 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours. Data expressed as mean ± SD. P < 0.05. (n=3).

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Figure 9. Alteration of intrahemocyte nitric oxide (NO) activity of juvenile S. serrata under exposure of 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours. Data expressed as mean ± standard deviation. P < 0.05. (n= 3).

Figure 10. Alteration of intrahemocyte phenoloxidase (PO) activity of juvenile S. serrata under exposure of 1, 2 and 3 ppm of sodium arsenite for 24, 48, 72 and 96 hours.. Data expressed as mean ± SD. P < 0.05. (n= 3).

Invertebrates including crustaceans lack of adaptive immune or nonlymphoid system and depend on innate immunity. The innate immune system is phylogenetically a more ancient defense mechanism than adaptive or lymphoid immunity. Crustacean host defence mechanisms do not show a high degree of specificity and lack adaptive immunological memory, but they

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have complex and efficient host defence systems that can identify and eliminate potential pathogens efficiently. Hemocytes or circulating cells of hemolymph are capable of responding to diverse forms of xenobiotics through aggregation, adhesion, phagocytosis and generation of cytotoxic molecules (Chakraborty et al., 2010; Ghosh, 2007; Saha et al., 2008; 2009 and 2010). The hemocyte density is considered as a useful indicator for other cell biological responses due to enhancement of the immune capability during stress leading to disease resistance (Truscott and White, 1990). Increment of THC indicates proliferation of hemocytes due to activation of hemocytopoietic organs of crab to release fresh hemocytes in blood stream (Victor, 1993). Hemocytes mediated cytotoxicity through production of reactive oxygen intermediates are in report in invertebrate (Holmblad and Soderhall, 1999; Nappi and Ottaviani, 2000). The superoxide anion is produced during reduction of molecular oxygen by capture of an electron and the reaction of nitric oxide, a degradation product of L- arginine which may generates peroxynitrite. Production of melanin through activation of phenoloxidase is identified as toxic elements to kill the invading pathogen (Manduzio et al., 2005; Arumugam et al., 2000; Tanner et al., 2006). Gradual decrease of aggregation, adhesion, phagocytosis of hemocytes and increment of intrahemocytes superoxide anion, nitric oxide and phenoloxidase were recorded against 1, 2 and 3 ppm of arsenite for 24, 48, 72 and 96 hours (Figure 8, 9 and 10). Situation may create a toxic environment within crab leading to possible impairment of immune status (Saha et al., 2007; 2008; 2009 and 2010). Sundarbans biosphere reserve of the state of West Bengal, India is a unique geographic area which supports diverse forms of marine and estuarine animals (Chakraborty, 1986). This ecological sensitive region of state receives a sustential load of silt and diverse xenobiotics carried by its extensive riverine system (Ghosh, 2007). Arsenic a major significant contaminant of this region is capable of initiating immunotoxicity in S. serrata, a commercially important bioresource (Gomez – Caminero et al., 2001). Sundarbans is reported to be under the threat of arsenic contamination (Das et al., 1995; Acharyya et al., 1999; Acharyya, 2002). Arsenic induced altered reactivity of hemocyte may render the juvenile mudcrab to become immunologically weak under parasitic attack. Such a situation may lead to a possible decline in population of this edible species in its natural habitat.

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ACKNOWLEDGMENT Authors thankfully acknowledge the Department of Science and Technology and University Grants Commission of Government of India for providing financial support through FIST and SAP – DRS - 1 for grant support, instrumentation and other facilities. The ethics committee of the department of Zoology of University of Calcutta is also gratefully acknowledged.

REFERENCES Acharyya, S. K., Chakraborty, P., Lahiri, S., Raymahashay, B. C., Guha, S. and Bhowmik, A. (1999). Arsenic poisoning in the Ganges delta. Nature, 401(6753): 545-550. Acharyya, S. K. (2002). Arsenic contamination in groundwater affecting major parts of southern West Bengal and parts of western Chattisgarh: Source and mobilization process. Curr. Sci., 82(6): 740-744. Ali, M. Y., Kamal, D., Hossain, S. M. M., Azam, M. A., Sabbir, W., Murshida, A., Ahmed, B. and Azam, K. (2004). Biological studies of the mud crab, Scylla serrata (Forskal) of the Sundarbans mangrove ecosystem in Khulna region of Bangladesh. Pakistan J. Biol. Sci., 7(11): 1981-1987. Anderson, R. S., Oliver, L. M. and Brubacher, L. L. (1992). Superoxide anion generation by Crassostrea virginica hemocytes as measured by nitroblue tetrazolium reduction. J. Invert. Pathol., 59: 303 – 307. Armstrong, P. B. (1980). Adhesion and spreading of Limulus blood cells to artificial surface. J. Cell Sci., 44: 243 – 262. Arumugan, M. Romestand, B. and Torreilles, J. (2000). Nitrite released in hemocytes from Mytilus galloprovincialis, Crassostrea gigas and Ruditapes decussatus upon stimulation with phorbol myristate acetate. Aquat. Living Resour., 13: 173 – 177. Aspan, A., Sturtevant, J., Smith, V. J. and Soderhall, K. (1990). Purification and characterization of a prophenoloxidase activating enzyme from crayfish blood cells. Insect Biochem., 20: 709 – 718. Bell, K. L. and Smith, V. J. (1993). In vitro super oxide production by hyaline cells of the shore crab, Carcinus maenas (L.). Dev. Comp. Immunol., 17: 211 – 219.

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Bodammer, J. E. (1978). Cytological observation on the blood and hemopoietic tissue in the crab, Callinectes sapidus. Cell Tiss. Res., 187: 79 – 96. Cerenius, L. and Soderhall, K. (2004). The prophenoloxidase activating system in invertebrates. Immunol. Rev., 198: 116 – 126. Chakraborty, S., Ray, M. and Ray, S. (2008). Sodium arsenite induced alteration of hemocyte density of Lamellidens marginalis - an edible mollusk from India. Clean, 36 (2): 195-200. Chakraborty, S., Ray, M. and Ray, S. (2010). Toxicity of sodium arsenite in the gill of an economically important mollusc of India. Fish and shellfish Immunol., 29: 136 – 148. Chakraborty, K. (1986). Fish and fisheries resources in the mangrove swamps of Sundarbans, West Bengal – an in depth study. Indian Forester, 112 (6): 538- 542. Chaudhuri, A. B. and Choudhury, A. (1994). Mangrove of Sundarbans Vol. I: India (Zakir Hussain, M. and Acharya, G. Ed.) IUCN, Bangkok, Thailand. The World Conservation Union. Chen, J. H. and Bayne, C. J. (1995). Bivalve (molluscs) hemocyte behaviours: characterization of hemocyte aggregation and adhesion and their inhibition in the California mussel (Mytilus californianus). Biol. Bull., 188: 255 – 266. Das, D., Chatterjee, A., Mandal, B. K., Samanta, G. and Chakraborti, D. (1995). Arsenic in ground water in six districts of West Bengal, India: the biggest arsenic calamity in the world. Part 2. Arsenic concentration in drinking water, hair, nails, urine, skin-scale and liver tissue (biopsy) of the affected people. Analyst, 120: 917 – 924. Das, S. and Roy Chowdhury, A. (2006). Arsenic: a global monster. Sci. Cult., 72(7- 8): 230 – 237. Ghosh, S. (2007). Immunological response of hemocytes of estuarine mud whelk exposed to diesel. M. Phil. Dissertation, University of Calcutta. Gomez - Caminero, A., Howe, P., Hughes, M., Kenyon, E., Lewis, D. R., Moor, M., Nag. J., Aitio, A. and Becking, G. (2001). Summary. In Environmental Health Criteria 224, Arsenic and Arsenic Compounds (Ed: 2nd), pp. 1-8: World Health Organization, Geneva. Green, L. C., Wagner, D. A., Glogowski, J., Skipper, P. L., Wishnok, J. S. and Tannenbaum, S. (1982). Analysis of nitrate, nitrite and (N15) nitrate in biological fluids. Anal. Biochem., 126: 131 – 138. Guria, K. and Ray, S. (2002). Modulation of glass surface adhesion characteristics and sub population shift of hemocytes of Bellamya

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bengalensis by synthetic pyrethroid. Indian J. Environ. Ecoplan., 6(1): 175-178. Heasman, M. P. and Fielder, D. R. (1983). Laboratory spawning and mass rearing of the mangrove crab (Scylla serrata Forskal), from first zoea to first crab stage. Aquacult., 34: 303-316. Holmblad, T. and Soderhall, K. (1999). Cell adhesion molecules and antioxidative enzymes in a crustacean, possible role in immunity. Aquacult., 172: 111 – 123. Hose, J. E., Marin, G. G. and Gerard, A. S. (1990). A decapod hemocyte classification scheme integrating morphology, cytochemistry and function. Biol. Bull., 178: 33 – 45. Johansson, M. W. and Soderhall, K. (1988). Isolation and purification of a cell adhesion factor from crayfish blood cells. J. Cell Biol., 106: 1795 – 1803. Kenney, D. M., Belamarich, F. A. and Shepro, D. (1972). Aggregation of horseshoe crab (Limulus polyphemus) amebocytes and reversible inhibition of aggregation by EDTA. Biol. Bull., 143: 548 – 567. Klassen, C. D. (1991). Principles of toxicology. In: Gilman, A. G., Tall, T. W., Nies, A. S., Taylor, P. (Eds.), Pharmacological Basis of Therapeutics, 8th ed. McGraw Hill, pp. 49 – 61. Krishnaja, A. P., Rege, M. S. and Joshi, A. G. (1987). Toxic effects of certain heavy metals (Hg, Cd, As, Se) on the intertidal crab Scylla serrata. Mar. Environ. Res., 21(2): 109 –119. Lesser, M. P. (2006). Oxidative stress in marine environment: biochemistry and physiological ecology. Annu. Rev. Physiol., 68: 253 – 278. Lowry, D. H., Rosebrough, M. J., Farr, A. L. and Randall, R. J. (1951). Protein measurement with the folinphenol reagent. J. Biol. Chem., 193: 265-275. Munoz, M., Cedeno, R., Rodriguez, J., Van der Knaap, W. P. W., Mialhe, E. and Bachere, E. (2000). Measurement of reactive oxygen intermediate production in hemocytes of the penaeid shrimp, Penaeus vannamei. Aquacult., 191: 89 – 107. Nappi, A. J. and Christensen, B. M. (2005). Melanogenesis and associated cytotoxic reactions: applications to insect innate immunity. Insect Biochem. Mol. Biol., 35: 443 – 459. Naskar, K. and Ghosh, A. (1989). Mangrove forest of the Sundarbans: Its impact on estuarine fisheries. Proceeding in Coast Zone Management of West Bengal, Sea Explores Institute, Calcutta, pp A 47 – 59.

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Ratcliffe, N. A. (1985). Invertebrate immunity – a primer for non-specialists. Immunol. Lett., 10: 253 – 270. Ravindranath, M. H. (1977). The circulating hemocyte population of the mole crab, Emerita asiatica Milne Edwards. Biol. Bull., 153(3): 415 – 423. Rodriguez, J. and Moullac, G. L. (2000). State of the art of immunological tools and health control of penaeid shrimp. Aquacult., 191: 109 – 119. Ray, S., Ray, M. and Saha, S. (2009). Aggregation and chemical induced interference of aggregation of hemocytes of Indian mud crab exposed to arsenic. Animal Biology Journal, 1(1): 27 – 37. Saha, S. and Ray, S. (2006). Hemocyte profile of the estuarine mud crab, Scylla serrata. Environ. Ecol., 24S (3A): 818 – 819. Saha, S., Ray, M. and Ray, S. (2007). Analyses of total count of hemocytes of estuarine crab, under acute arsenic exposure. Icfai J. Life Sci., 1(3): 75 – 78. Saha, S., Ray, M. and Ray, S. (2008). Kinetics of nonself surface adhesion and phagocytic response of hemocytes of Scylla serrata exposed to sodium arsenite. Toxicol. Int., 15(1): 15 – 19. Saha, S., Ray, M. and Ray, S. (2008). Nitric oxide generation by immunocytes of mud crab exposed to sodium arsenite. In Zoological Research in Human Welfare, Zoological Survey of India, Kolkata Publication, paper 44: 425 – 428. Saha, S., Ray, M. and Ray, S. (2009). Activity of Phosphatase in the Hemocytes of Estuarine edible mud crab, Scylla serrata Exposed to Arsenic. Journal of Environmental Biology, 30(5): 655 – 658. Saha, S., Ray, M. and Ray, S. (2009). Effect of sublethal concentration of arsenic on hemocyte density in edible crab, Scylla serrata. Animal Biology Journal. 1(1): 17 – 26. Saha, S., Ray, M. and Ray, S. (2009). Recognition of antilymphocyte and antihemocyte sera by crab (Scylla serrata) hemocytes exposed to arsenic. Research in Environment and Life Science, 2(1): 1 – 6. Saha, S., Ray, M. and Ray, S. (2010). Shift in cytoarchitecture of immunocytes of mudcrab exposed to arsenic. International Journal of Applied Biology and pharmaceutical Technology, 1(2): 234 – 246. Saha, S., Ray, M. and Ray, S. (2010). Screening of phagocytic response and intrahemocytotoxicity of Scylla serrata under the challenge of charcoal particle exposed to arsenic. Asian Journal of Experimental Biological Sciences. 1(1): 47 – 54.

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Saha, S., Ray, M. and Ray, S. (2010). Behavioural shift of estuarine mudcrab as biomarker of arsenic exposure in Sundarbans estuary of West Bengal. Journal of Applied and Natural Science. 2(2): 258 – 262. Sahoo, B., Sethi, S. Mishra, B.K. and Das, B. K. (2005). Effects electors on prophenoloxidase and superoxide anion activities of freshwater prawn, Macrobrachium malcolmsonii. Asia Fisheries Sci., 18: 345 – 353. Sarkar, S. K., Bhattacharya, B., Debnath, S. Bandopadhaya, G. and Giri, S. (2002). Heavy metals in biota from Sundarbans wetland ecosystem, India: Implications to monitoring and environmental assessment. Aquatic Ecosystem Health and Management, 5(4): 467 – 472. Sarkar, S. K., Franciskovic – Bilinski, S., Bhattacharya, M., Saha, M. and Bilinski, H. (2004). Level of elements in the surficial estuarine sediments of the Hugli River, northeast India and their environmental implications. Environ. Int., 30 (8): 1089 – 1098. Sauve, S., Brousseau, P., Pellerin, J., Morin, Y., Seneeal, L., Goudreau, P. and Fournier, M. (2002). Phagocytic activity of marine and freshwater bivalves: in vitro exposure of hemocytes to metals (Ag, Cd, Hg and Zn). Aquat. Toxicol., 58: 189 – 200.Rodriguez and Moullac, 2000; Shibu Vardhannan, Y. and Radhakrishnan, T. (2002). Acute toxicity of copper, arsenic and HCH to paddy field crab, Paratelphusa hydrodromus (Herb.). J. Environ. Biol., 23(4): 387 – 392. Soderhall, K. and Cerenius, L. (1998). Role of the prophenoloxidaseactivating system in invertebrate immunity. Curr. Opin. Immunol., 10: 23 – 28. Soderhall, K., Smith, V. J. (1983). Seperation of the hemocyte populations of Carcinus maenus and other marine decapods, and proPO distribution. Dev. Comp. Immunol., 7: 229-239. Song, Y. L. and Hsieh, Y. T. (1994). Immunostimulation of tiger shrimp (Penaeus monodon) hemocytes for generation of microbicidal substances. Analysis of reactive oxygen species. Dev. Comp. Immunol., 18: 201 – 209. Sritunyalucksana, K. and Soderhall, K. (2000). The proPO and clotting system in crustaceans. Aquacult., 191: 53 – 69. Sung, H. H., Chang, H. J. Her, C. H., Chang, J. C. and Song, Y. L. (1998). Phenoloxidase activity of hemocytes derived from Penaeus monodon and Macrobrachium rosenbergii. J. Invert. Pathol., 71(1): 26 – 33. Takahashi, H., Azumi, K. and Yokosawa, H. (1994). Hemocyte aggregation in the solitary ascidian Holocynthia roretzi : Plasma factors, magnesium

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ion and Met-Lys-Bradykinin induce the aggregation. Biol. Bull., 186: 247 – 253. Takahashi, H., Azumi, K. and Yokosawa, H. (1995). A novel membrane glycoprotein involved in ascidian hemocyte aggregation and phagocytosis. Eur. J. Biochem., 233: 778 – 783. Tanner, C. A., Burnett, L. E. and Burnett, K. G. (2006). The effects of hypoxia and pH on phenoloxidase activity in the atlantic blue crab, Callinectes sapidus. Comp. Biochem. Physiol., A144: 218 – 223. Truscott, R. and White, K. N. (1990). The influence of metal and temperature stress on the immune system of crabs. Funct. Ecol., 4: 455 –461. Victor, B. (1993). Responses of hemocytes and gill tissues to sublethal cadmium chloride poisoning in the crab, Paratelphusa hydrodromous (Herbst.). Arch. Environ. Contam. Toxicol., 24: 432 – 439.

In: Aquatic Animals Editor: David L. Eder

ISBN: 978-1-61470-123-1 © 2012 Nova Science Publishers, Inc.

Chapter 3

THE POTENTIAL THREAT OF GENOTOXIC METALS TO MARINE MAMMAL HEALTH: A CASE STUDY OF CHROMIUM TOXICITY IN TOOTHED AND BALEEN WHALES Carolyne LaCerte and John Pierce Wise, Sr.* Wise Laboratory of Environmental and Genetic Toxicology, Maine Center for Toxicology and Environmental Health, University of Southern Maine, Portland, Maine 04103, U.S.

ABSTRACT Ocean pollution has emerged as one of the greatest threats to marine animal heath with recent data showing that metal pollution has reached even the remotest ocean regions. Marine mammals are at particular risk as they integrate all possible routes of exposure: inhalation, ingestion and dermal pathways. Many marine mammals have become critically endangered and are failing to recover. Marine mammals are important species for ecosystem health and for the economies of coastal communities. Moreover, they are charismatic species that capture people's attention and remind them of the

* Correspondence: Dr. John Pierce Wise, Sr.; Address: 96 Falmouth St. PO Box 9300, Portland, ME 04104-9300; Tel: (207) 228-8050. Fax: (207) 228-8518. Email: [email protected]

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Carolyne LaCerte and John Pierce Wise, Sr. importance of the ocean environment. Understanding them and the threats that face them is a key need in protecting the ocean environment. We have been pioneering the study of metals, particularly genotoxic metals as particular threat to marine mammal health. Marine mammals are exposed to metals and if these metals can damage DNA in these animals, the consequences could impact both the individual, through metal-induced disease, and the population as a whole, by impairing reproduction. In this chapter, we focus on chromium as a representative genotoxic metal and describe the levels of Cr in tissues from free ranging, healthy whales considering both toothed and baleen whales. We then determine the genotoxic effects of Cr in cultured whale cells and compare genotoxic doses to tissue levels. The chapter discusses the potential threat of genotoxic metals, like chromium, what the sources of exposure are in the marine environment, and compares the data to effects seen in humans and laboratory animals.

INTRODUCTION Marine pollution poses a significant ecological and health problem. The oceans cover approximately 70% of the Earth‘s surface thus influencing numerous biological and physical processes on the planet. The oceans are polluted with toxic substances from both natural and anthropogenic sources. Industrial pollution and oil spills have released pollutants into the oceans at high concentrations over short time periods, which inhibits the evolutionary ability of marine organisms to adapt and thus poses a significant threat to marine life and to humans who rely on them [1]. Some marine populations are in crisis and ocean pollution may be playing a significant role. For example, the populations of North Atlantic right whales (Eubaelena glacialis) and Western Alaskan Steller sea lions (Eumetopias jubatus) have declined with limited recovery. The North Atlantic right whale population is estimated to be about 400 individual whales. The main causes of their mortality are collisions with ships and entanglements in fishing gear [2]; however these whales also have a critical reproductive problem, which may be caused by or exacerbated by exposure to environmental contaminants [3]. The once abundant Western Steller sea lion population suffered a 75% population reduction from 1979 to 1997 [4] that still continues at a rate of 5.2% per year [5]. The specific cause of this decline is unknown but is considered to be environmental and pollution is

The Potential Threat of Genotoxic Metals to Marine Mammal Health 79 one of the likely factors [6]. More studies are needed to fully understand the impact of pollution on marine life. One significant threat that ocean pollution poses to marine species is the potential to damage their DNA. All life is dependent on its DNA for reproduction and survival. DNA damage can cause mutations, alter or delete chromosomal structures, and affect gene expression. These effects can manifest as various negative health consequences including reproductive failure, developmental abnormalities, cancer, neurotoxicity, immunotoxicity among others. Although, DNA damage has not been studied in a wild marine population, it certainly could be contributing to the distress of marine populations such as the right whale and Steller sea lion. There is a wide spectrum of pollutants that might cause genotoxicity in marine species including organic chemicals like polycyclic aromatic hydrocarbons (PAHs) and metals. One environmental pollutant of emerging marine health concern is the metal chromium (Cr). Estimates indicate that 33 tons of Cr are released into the environmental annually [7]. In the marine environment, the hexavalent form of Cr [Cr(V)] is substantially more prominent than the trivalent form Cr(III). Cr(VI) damages DNA causing chromosomal aberrations [8,9], DNA double strand breaks [10] and oxidative damage [11]. Consistent with its genotoxicity, Cr(VI) causes cancer and reproductive effects [12]. Given these outcomes, we decided to evaluate Cr as a pollutant of concern for marine life. We decided to use whales as model species because they have long lifespans, occupy top positions in the trophic system and best represent humans in the environment. In particular, we chose two species; the North Atlantic right whale, a baleen whale, and the sperm whale (Physeter macrocephalus), a toothed whale. We chose the North Atlantic right whale because of its known reproductive problems, and the proximity of its range to urban human centers (i.e. Eastern seaboard of the United States). By contrast, the sperm whale has a global distribution and is usually further away from land, and as a toothed whale, it feeds higher on the food chain consuming fish and squid, while right whales are baleen feeders eating mostly tiny copepods. The frank challenge of studying the impacts of pollution on marine species is the lack of experimental models that can be controlled in the laboratory. The traditional experiment would be to study several groups of animals in the laboratory after exposing them to increasing doses of the pollutant of interest. For marine mammals, this type of experimentation is technically impractical (e.g. difficult to have multiple tanks of whales to

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dose) and ethically difficult (e.g. should one be dosing endangered animals with toxic chemicals). A similar challenge exists for humans. Thus, toxicologists have employed rodent studies to address some of these questions. However, rodents are known to be useful but often-imprecise models for humans and most likely this outcome is also true for marine species. To gain species specific information about people, human cell lines were developed and a variety of molecular and cellular tests have been developed to test toxicity. Cell culture studies can not only provide species specific information, but also differences in cell type and cell origins (i.e. where the cells were derived from). In addition, very specific cellular investigations can be performed when using cell cultures including but not limited to intracellular activity (e.g. replication and transcription of DNA), intracellular flux (e.g. membrane trafficking, signal transduction), environmental interaction (e.g. cytotoxicity, carcinogenesis), cell to cell interaction (e.g. cell proliferation, cell adhesion and motility), and genetics (e.g. genome analysis), which cannot be done using whole animal models. Thus, marine cell lines provide a valuable tool to study the impact of pollutants on marine species. Accordingly, this chapter described an approach to evaluating the impact of marine pollution focusing on Cr(VI) and using cell lines from sperm whales and North Atlantic right whales.

MATERIALS AND METHODS Biopsy Collection Techniques for the collection of skin biopsy samples from cetaceans are well established [13]. Biopsy samples that were taken from both right whales and sperm whales were collected with a stainless steel cylindrical biopsy tip that was 7 mm in diameter, and contained a flare rim and stop collar to prevent penetration beyond the thickness of the skin and blubber interface. Biopsy tips were attached to an arrow and loaded into either a crossbow or compound longbow depending on the preference of the person using the equipment. To ensure that there was no contamination by the biopsy dart or the forceps used to extract the biopsy, a piece of whale tissue provided by Inupiat hunters (during their annual fall hunt in Barrow‘ Alaska) was sampled by the same process with forceps and the biopsy dart and the sample comparison was tested with a Teflon knife. There was no difference in the

The Potential Threat of Genotoxic Metals to Marine Mammal Health 81 levels of chromium between the two samples and chromium levels were low indicating that there was no contamination by the forceps and darts. In addition, the dart and forceps were soaked in 2% HNO3 solution in order to measure chromium leachate and found that the levels of chromium were nondetectable.

Development of Whale Cell Lines North Atlantic right whale skin, lung and testes fibroblasts were isolated from tissue explants obtained during necropsy of an animal that died while in captivity. Sperm whale skin fibroblasts were isolated from a tissue biopsy obtained from a free-ranging whale off the coast of North Carolina. All North Atlantic right whale cells were from the same individual whale. Each tissue type was handled independently in a different biosafety hood in order to avoid cross-contamination. Tissue explants were rinsed several times in phosphate-buffered saline with penicillin-streptomycin and gentamicin. Explants were then segmented into less than or equal to 1 mm pieces with a scalpel, rinsed repeatedly and placed in completed culture medium that consists of DMEM/F12 that is supplemented with 15% cosmic calf serum, 2 mM L-glutamine, 100 U/ml penicillin, 100 ug/ml streptomycin and 0.1 mM sodium pyruvate, and placed in a 33 °C humidified incubator with 5% CO2. Cells were examined for growth out of each tissue explant. In this case study, we focused on somatic fibroblasts that have originated from lung, testes, and skin tissues.

Cr(VI) Compounds Two representative hexavalent chromium compounds were used in this case study. Lead chromate was used as a representative particulate Cr(VI) compound and treatments were dispensed from a suspension of particles to ensure cells were exposed to intact particles. Sodium chromate was used as a representative soluble Cr(VI) compound model and was administered as a solution in water. Lead chromate treatments were expressed as weight per surface area (ug/cm2) and sodium chromate concentrations as uM.

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Inductively Coupled Plasma Mass Spectroscopy Whale skin samples were analyzed for total Cr using inductively coupled plasma mass spectrometry (ICPMS) according to our published methods using a Perkin Elmer/ Sciex ELAN ICPMS [14]. Interference check solutions were analyzed with all sample runs to compensate for any matrix effects which might be interfering with sample analysis. Standard quality assurance procedures were employed. Instrument response was evaluated initially, after every 10 samples, as well as at the end of each analytical run using a calibration verification standard and blank. All data are presented as ug total Cr/g tissue wet weight.

Measuring Cell Death In order to investigate the effect of Cr(VI) on the survival and proliferation of cetacean cells we used a clonogenic cytotoxicity assay. Specifically it measures the reduction in plating efficiency in treatment groups relative to the controls. Briefly, 100,000 cetacean cells in log phase growth were seeded into each of a 6-well tissue-culture dish with 2.3 ml of medium into each well. Cells were allowed to grow for 48 h before treatment. Cells were then treated for 24 h with either lead chromate or sodium chromate. After treatment, medium was collected in order to include any loosely adherent mitotic cells, rinsed with phosphate-buffered saline solution and detached from the dish with trypsin. Cells were then re-suspended in fresh medium and reseeded at 1,000 cells per 100 mm dish with four dishes per treatment group. Colonies formed in about 14 days and were fixed and stained with crystal violet. Colonies were then counted in each dish and averaged together to get a mean value for each treatment in each experiment. Each experiment was done independently and repeated at least three times.

Measuring Clastogenicity In order to investigate if Cr(VI) was genotoxic to cetacean DNA, clastogenicity was determined by measuring the amount of chromosomal damage in treatment groups and controls. Cells were prepared for chromosome analysis by seeding 800,000 cells in each 100 mm dish and allowed to grow for 48 h. The cultures were treated for 24 h with either a

The Potential Threat of Genotoxic Metals to Marine Mammal Health 83 suspension of lead chromate or with sodium chromate. Demecolchicine at 0.1 ug/ml was added 5 hr before the end of treatment time in order to arrest the cells in metaphase. Cells were collected and resuspended in 10 ml of 0.075 M postassium chloride hypotonic solution for 17 minutes followed by fixation in methanol:acetic acid (3:1). The fixative was changed twice and the cells were dropped on to clean wet slides and uniformly stained using a 5% Giemsa stain in Gurr‘s buffer. At least three independent experiments were conducted. Two measures were considered, the percent of metaphases with damage and total damage. The percent of metaphases with damage measured the frequency that a metaphase cell incurs at least one damaging event. For this measure, whether a metaphase had 1 or 5 broken chromosomes, it was recorded as 1 metaphase with chromosome damage. By contrast, total damage in 100 metaphases with damage measured multiple aberrations within one metaphase cell. For this measure, if a metaphase had 1 or 5 broken chromosomes, it was recorded as 1 or 5 aberrations in that cell.

RESULTS Determining Cr in Levels in Whale Skin The first step in assessing the impact of marine pollution on a marine species is to determine the level of a toxic agent in the tissue of the marine species. For our Cr case study, we measured the levels of total Cr in skin biopsies from seven free-ranging North Atlantic right whales from the Bay of Fundy, and in skin biopsies from 361 free-ranging sperm whales from around the world. Cr was present in all but two sperm whales. For Cr, it cannot be determined which valence state of Cr the whales were originally exposed to because in mammalian tissues and systems Cr(VI) is rapidly metabolized to Cr(III) after exposure [15]. However, one can assess the total Cr exposure. We found total Cr levels in right whales ranged from 4.9 to 10 ug Cr/g tissue with a mean of 7 ug/g w.w., and total Cr levels in sperm whales ranged from 0.9 to 122.6 ug Cr/g tissue with a mean of 8.8 ug/g w.w. In the sperm whale with global distributions the highest levels were found in sperm whales sampled near the islands of Kiribati in the Pacific Ocean and in the Seychelles in the Indian Ocean.

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Evaluating the Toxicity of Cr to Cetaceans Having determined tissue levels, the next step was to address whether these levels were low or high and whether they might impact whale health. One cannot do the obvious experiment and expose whales directly to Cr and determine its toxicity and potency. Instead, we have to use other measures. Thus, to provide this context we turn to the published literature and whale cell culture based studies.

Toxicity Context from the Cr Literature In considering the published literature, to assess if the levels we see in whales are of concern, we need to consider studies that measure tissue Cr levels in the context of toxicity. We did not find any laboratory animal-based studies of Cr tissue levels in the context of toxicity. Some rodent studies determined Cr distribution in the body, but they did not attribute those levels to any toxic outcome. In humans, we found studies of Cr levels in workers with Cr(VI)-induced lung cancer. Like the whale skin, these levels were measured as total Cr. The levels ranged from 0.4-132 ug/g lung tissue with a median level of 20.4 ug/g in their lungs [16]. Cr was measured in the skin of one of these workers and was 0.05 ug Cr/g [17]. These data suggest the whale levels are at a level of concern and therefore, should be considered high, perhaps very high. Toxicity Context from Whale Cell Culture Studies Cell culture studies have the advantage of providing species specific information in an experimentally controlled setting. They are of course limited to cellular and molecular outcomes, but they are valuable tools for species that are difficult to study in the laboratory, such as whales and humans. In this case study, we consider two measures of cellular and molecular toxicity, cytotoxicity, which integrates all cellular events that lead to cell death, and genotoxicity, which reflects the events that cause DNA damage. In this study, we consider both sperm whale and North Atlantic right whale cells. Cr Cytotoxicity Studies in Whale Cells Cytotoxicity studies provide two types of context. First, they indicate concentrations at which cell death occurs, which, if it occurs in vivo, can be detrimental to the tissue or organism. Second, they provide context to other cell culture studies indicating which concentrations are of low, medium, high

The Potential Threat of Genotoxic Metals to Marine Mammal Health 85 and supra toxicity. Thus, if another effect is observed at a concentration that cells survive, one is concerned it might occur in vivo. On the other hand, if the effect only occurs at supratoxic concentrations, then it is unlikely to occur in vivo because the cell would likely have died before it could manifest itself. We investigated the cytotoxicity of both soluble and particulate Cr(VI) in cell lines from both whales. Soluble Cr(VI) induced cytotoxicity in a concentration-dependent manner in all whale cell lines that we tested (Figure 1). In right whale lung fibroblasts, concentrations of 1, 2.5, 5, 10 and 25 uM sodium chromate induced 88, 74, 56, 32 and 0% relative survival, respectively (Figure 1A). These concentrations induced 85, 63, 23, 2 and 0% relative survival, respectively, in right whale testes fibroblasts (Figure 1A), and 78, 60, 43, 10 and 0% relative survival, respectively, in right whale skin fibroblasts (Figure 1A and 1B). In sperm whale skin fibroblasts, at the same concentrations soluble Cr(VI) induced 77, 58, 20, 4 and 0% relative survival, respectively (Figure 1B). Particulate Cr(VI) also induced cytotoxicity in a concentration-dependent manner in all of the whale cell lines that we tested (Figure 2). Concentrations of 0.1, 0.5, 1, 5, and 10 ug/cm2 induced 108, 100, 78, 14 and 1% relative survival, respectively, in right lung fibroblasts (Figure 2A) and 126, 64, 49, 11, and 1% relative survival, respectively (Figure 2A and 2B), in right whale skin fibroblasts. In sperm whale skin fibroblasts, the same concentrations of particulate Cr(VI) induced 91, 68, 63, 36, and 7% relative survival, respectively (Figure 2B). Right whale testes cells were not tested for particulate Cr(VI)-induced cytotoxicity, because we reasoned a Cr(VI) particle was unlikely to cross the blood-testis barrier. The data show that Cr(VI) is cytotoxic to whales cells and thus, will likely be cytotoxic in vivo at certain doses. The data also show that cells from different organs can have different responses as right whale skin cells were more sensitive to Cr(VI) than right whale lung cells and right whale testes cells were even more sensitive than the skin cells. The data further show that species differ in their response as sperm whale skin cells were more sensitive to soluble Cr(VI) than right whale skin cells and less sensitive than to particulate Cr(VI).

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Figure 1. Cytotoxicity of Soluble Cr(VI) in North Atlantic Right Whale and Sperm Whales Cells. Soluble Cr(VI) is cytotoxic to both cetacean species, in all cell types tested in a concentration-dependent manner after a 24 h exposure. A) North Atlantic right whale lung, testes, and skin fibroblasts. B) North Atlantic right whale skin fibroblasts and sperm whale skin fibroblasts.

The Potential Threat of Genotoxic Metals to Marine Mammal Health 87

Figure 2. Cytotoxicity of Particulate Cr(VI) in North Atlantic Right Whale and Sperm Whales Cells. Particulate Cr(VI) is cytotoxic to both cetacean species, in all cell types tested in a concentration-dependent manner after a 24 h exposure. A) North Atlantic right whale lung and skin fibroblast cells. B) North Atlantic right whale skin fibroblasts and sperm whale skin fibroblasts.

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Cr Genotoxicity Studies in Whale Cells A critical part of assessing the risk posed by a chemical is to understand its genotoxic potential as genotoxic chemicals are likely to be carcinogenic and interfere with embryogenesis and development. For this case study, we present chromosomal damage as a measure of genotoxic damage because chromosomal rearrangements and alterations are commonly found in cancers and reproductive and development disorders; are a standard hazard characterization for the U.S. Environmental Protection Agency (EPA) and the Food and Drug Administration (FDA); and are often cited in the workplace health standards for the Occupational Safety and Health Administration (OSHA). We investigated the genotoxicity of both a soluble and particulate form of Cr(VI) in whale cell lines. Soluble sodium chromate induced chromosome damage in a concentration-dependent manner in all whale cell lines that we tested (Figures 3, 4 and 5). In right whale lung fibroblasts, concentrations of 1, 2.5, 5, and 10 uM soluble Cr(VI) damaged chromosomes in 15, 20, 24, and 49% of metaphases and induced 15, 21, 27, and 63 aberrations per 100 metaphases, respectively (Figure 3A and 3A). In right whale skin fibroblasts, the same concentrations of sodium chromate damaged chromosomes in 7, 15, 23 and 33% of metaphases and induced 8, 18, 30 and 48 aberrations per 100 metaphases respectively (Figure 3B and 3B). In right whale testes fibroblasts, the same concentrations damaged chromosomes in 11, 18, 32 and 41% of metaphases and induced 12, 22, 39 and 58 aberrations per 100 metaphases, Figure 3C). In sperm whale skin fibroblasts, concentrations of 1, 2.5, 5 and 10 uM soluble Cr(VI) damaged chromosomes in 7, 16, 20 and 31% metaphases and induced 7, 16, 24 and 40 aberrations per 100 metaphases, respectively (Figure 3B and 3B). There was no difference in soluble Cr(VI)induced chromosome damage between right whale and sperm whale skin fibroblasts.

The Potential Threat of Genotoxic Metals to Marine Mammal Health 89

Figure 3. Genotoxicity of Soluble Cr(VI) in North Atlantic Right Whale and Sperm Whale Cells. Soluble Cr(VI) is clastogenic to North Atlantic right whale lung, skin, testes fibroblast cells and sperm whale skin fibroblast cells in a concentrationdependent manner after a 24 h exposure measured as the percent of metaphases with chromosome damage. To allow for easier comparisons background levels of chromosome damage were subtracted. A) North Atlantic right whale lung, testes and skin fibroblasts. Background levels were 7, 6, and 5 percent of metaphases with chromosome damage, respectively. B) North Atlantic right whale skin fibroblasts and sperm whale skin fibroblasts. Background levels were 5 percent of metaphases with chromosome damage for both species.

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Figure 4. Genotoxicity of Soluble Cr(VI) in North Atlantic Right Whale and Sperm Whale Cells. Soluble Cr(VI) is clastogenic to North Atlantic right whale lung, skin, testes fibroblasts and sperm whale skin fibroblasts in a concentration-dependent manner after a 24 h exposure measured as the total damage in 100 metaphases. To allow for easier comparisons background levels of chromosome damage were subtracted. A) North Atlantic right whale lung, testes and skin fibroblasts. Background levels were 7, 6, and 5 total chromosome damage in 100 metaphases, respectively. B) North Atlantic right whale skin fibroblasts and sperm whale skin fibroblasts. Background levels were 5 total chromosome damage in 100 metaphases for both species.

The Potential Threat of Genotoxic Metals to Marine Mammal Health 91

Figure 5. Genotoxicity of Particulate Cr(VI) to North Atlantic Right Whale and Sperm Whale Cells. Particulate Cr(VI) is clastogenic to North Atlantic right whale lung and skin fibroblasts and sperm whale skin fibroblasts in a concentrationdependent manner after a 24 h exposure measured as the percent of metaphases with chromosome damage. To allow for easier comparisons background levels of chromosome damage were subtracted. A) North Atlantic right whale lung and skin fibroblasts. Background levels were 7 and 9 percent of metaphases with chromosome damage respectively. B) Particulate Cr(VI) is genotoxic to North Atlantic right whale skin fibroblasts and sperm whale skin fibroblasts. Background levels were 9 and 2 percent of metaphases with chromosome damage respectively.

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Figure 6. Genotoxicity of Particulate Cr(VI) to North Atlantic Right Whale and Sperm Whale Cells. Particulate Cr(VI) is clastogenic to North Atlantic right whale lung and skin fibroblasts and sperm whale skin fibroblasts in a concentrationdependent manner after a 24 h exposure measured as the total damage in 100 metaphases. To allow for easier comparisons background levels of chromosome damage were subtracted. A) North Atlantic right whale lung and skin fibroblasts. Background levels were 8 and 10 total chromosome damage respectively. B) North Atlantic right whale skin fibroblasts and sperm whale skin fibroblasts. Background levels were 10 and 2 total chromosome damage respectively.

The Potential Threat of Genotoxic Metals to Marine Mammal Health 93 Particulate Cr(VI) induced chromosome damage in a concentrationdependent manner in all whale cell lines that we tested (Figures 5 and 6). In ight whale lung fibroblasts, concentrations of 0.5, 1 and 5 ug/cm2 lead chromate damaged chromosomes in 14, 15 and 17% of metaphases and induced 17, 17, and 20 aberrations per 100 metaphases respectively (Figure 5A and 6A). In right whale skin fibroblasts, the same concentrations of lead chromate damaged chromosomes in 16, 19 and 26% of metaphases and induced 17, 22 and 30 aberrations per 100 metaphases, respectively (Figure 5B and 6B). In sperm whale skin fibroblasts, the same concentrations damaged chromosomes in 6, 12 and 27% of metaphases and induced 7, 13 and 28 aberrations per 100 metaphases, respectively (Figure 5B and 6B).

CONCLUSION Cr is released into the marine environment by both natural and anthropogenic sources with the primary anthropogenic sources including fuel combustion, wastewater effluent, leather tanning industries and textile manufacturing and the natural source being continental dust flux. In this Cr case study the Cr levels found in whales show that indeed Cr is a global pollutant. Both the literature in occupationally exposed humans and in studies of cultured whales cells indicates that these levels are high and likely to induce DNA damage and cell death in sufficiently exposed whales. The challenge of course is how to extrapolate from the cell culture doses to levels in the animals themselves. Converting the experimentally administered doses in cell culture to parts per million (ppm) allows for direct comparison of the administered doses to the tissue levels (Table 1). For example, the lowest genotoxic doses we measured for Cr(VI) converts to 0.052 ppm (1 uM sodium chromate). This value is 135-times lower than the mean level found in right whales (7 ppm). It is 169-times lower than the mean level found in sperm whales (8.8 ppm) and 2,358-times lower than the highest level found in sperm whales (122.6 pm). It should also be noted that we did not seek to find the lowest level that causes DNA damage nor did we explore longer exposure times, thus the potential is that damage can occur at concentrations even lower than these.

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Table 1. Soluble and Particulate Cr(VI) Treatments Converted to Parts Per Million (ppm)

uM 1 2.5 5 10 25

Sodium Chromate* ppm 0.052 0.13 0.26 0.52 1.3

ug/cm2 0.1 0.5 1 5 10

Lead Chromate** ppm 0.068 0.340 0.681 3.404 6.807

*Calculation used for sodium chromate: ([Cr] x MW(Cr))/1000 = ppm; Eg. (1 x 52)/1000 = 0.052 **Calculation used for lead chromate: [([Cr] x [cell culture dish area] x MW(Cr))/ MW(lead chromate)]/ media volume = ppm; Eg. [(0.1 x 55 x 52)/323.19]/13 = 0.068

An alternative approach would be to convert the tissue levels to Cr concentrations (Table 2). The lowest sperm whale level detected was 0.9 ppm and the highest was 122.6 ppm with a mean of 8.8 ppm. These values convert to 17 uM, 2,357 uM and 169 uM respectively. In right whales, the lowest level was 4.9 ppm the highest was 10 ppm and the mean was 7.1 ppm. These values convert to 95, uM, 201 uM and 137 uM respectively. We found genotoxicity and cytotoxicity beginning at 1 uM Cr(VI), though we did not test lower concentrations to determine the lowest genotoxic dose. Thus, the values in whale tissues are 17-2,357 times higher. Table 2. Parts per Million (ppm) of Chromium Converted to Micro Molar (um) North Atlantic Right Whale

Sperm Whale

ppm

uM

ppm

uM

Minimum Detected

4.9

94.8

0.9

17.31

Maximum Detected

10

201.1

122.6

2357.69

Global Mean

7.1

137

8.8

169.23

The Potential Threat of Genotoxic Metals to Marine Mammal Health 95 Of course, we cannot know how much of the Cr was Cr(VI) and how much was Cr(III). Cr(III) toxicity is unknown in whale cells, though at high doses such as the higher end of our range, it can cause toxicity and damage DNA [18]. Similarly, we cannot know the exposure pattern to Cr in the whales- were they exposed to a lot of Cr in a short time or less Cr over a longer time. However, data in human cells show that exposure to lesser amounts of Cr over a longer period of time also cause damage. The data show that whales do accumulate high Cr levels and Cr is genotoxic to whale cells at levels that are much lower than those they accumulate. Thus, it seems reasonable to conclude that Cr is a genotoxic threat to whales with more work needed to determine the exposure levels needed to cause this effect. Cr(VI) induced chromosome damage in both baleen and toothed whales in multiple organs. In humans, this outcome is thought to underlie the carcinogenicity of Cr(VI). It is unknown if this is a concern for whales. Cancer does occur in whales. Liver hemangiomas, genital papillomas, and skin and jaw fibromas were found in sperm whale necropsies, with no known cause, though Cr was not considered [19]. Previous studies have linked PAHs to the cause of cancer found in the St. Lawrence Estuary beluga whales [20] though again Cr was not considered. Thus, it remains unknown if Cr(VI) is contributing to a cancer burden in whales; however, our data suggest that it could be a contributing factor. In addition, in our approach we found that Cr(VI) was cytotoxic and genotoxic to the North Atlantic right whale testes cells. This outcome would suggest that Cr(VI) could cause reproductive problems in whales if it reached sufficient levels in the testes. Such a possibility is consistent with the fact that right whales are exposed to high levels of Cr and the lack of the population‘s ability to rebound. Human and rodent studies have shown that Cr(VI) accumulates in the testes and causes a reduction in testicular weights, seminiferous tubule degeneration, decreased sperm counts and altered reproductive behaviors [17, 21-25]. We then hypothesize that heavy metal exposure may be contributing to the North Atlantic right whales problems with reproductive fitness and population recovery, specifically Cr(VI). More work is needed to investigate this hypothesis. By migrating within 50 miles of urbanized coastal areas we have previously calculated the potential exposure scenario in a 24 hour period to the North Atlantic right whales based upon whale lung volume, respiration rate and atmospheric Cr levels, to be 260 ug of Cr [14]. However, our calculations may be an understatement because it has previously been suggested that coastal marine mammals might be more at risk from

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atmospheric pollution than humans due to the concentration of particles in the air-water interface [26]. Therefore, Cr is an environmental pollutant of concern for both humans and marine mammals.

REFERENCES [1]

Tierney, KB; Kennedy, CJ. Background Toxicology. In: Editor PW, Editor SS, Editor LF, Editor HS, Editor WG. Oceans and Human Health Risks and Remedies from the Seas. Elsevier; 2008;145. [2] Kraus, SD; Brown, MW. North Atlantic right whales in crisis. Science, 2005, 309, 561-562. [3] Kraus, SD; Hamilton, PK. Reproductive parameters of the North Atlantic right whale. J Cetacean Res Manag (special issue), 2001, 2, 213-236. [4] Calkins, DG; Mallister, DC. Steller sea lion status and trend in Southeast Alaska: 1979-1997. Mar Mammal Sci, 1999, 15, 462-477. [5] Loughlin, TR; York, AE. An accounting of the sources of Steller sea lion, Eumetopias jubatus, mortality. Mar Fish Rev, 2000, 62, 40-45. [6] Barron, MG; Heintz, R. Contaminant exposure and effects in pinnipeds: implications for Steller sea lion declines in Alaska. Sci Total Environ. 2003, 311, 111-133. [7] Agency for Toxic Substances and Disease Registry (ATSDR) 2008. Toxicological profile for Chromium. Atlanta, GA: U.S. Department of Health and Human Services, Public Health Service. [8] Wise, JP; Wise, SS. The cytotoxicity and genotoxicity of particulate and soluble hexavalent chromium in human lung cells. Mutat Res. 2002, 517, 221-229. [9] Wise, SS; Holmes, AL. Chromium is the proximate clastogenic species for lead chromate-induced clastogenicity in human bronchial cells. Mutat Res. 2004, 560, 79-89. [10] Xie, H; Wise, SS. Carcinogenic lead chromate induces DNA doublestrand breaks in human lung cells. Mutat Res. 2005, 586, 160-172. [11] Bagchi, D; Bagchi, M. Chromium (VI)-induced oxidative stress, apoptotic cell death and modulation of p53 tumor suppressor gene. Mol Cell Biochem. 2001, 222, 149-158. [12] Costa, M. Toxicity and carcinogenicity of Cr(VI) in animal models and humans. Crit Rev Toxicol. 1997, 27, 431-442.

The Potential Threat of Genotoxic Metals to Marine Mammal Health 97 [13] Brown, MW; Kraus, SD. Reaction of right whales (Eubalaena glacialis) to skin biopsy sampling for genetic and pollutant analysis. Reports Intern Whaling Commission (special issue), 1991, 13, 81-89. [14] Wise, JP; Wise, SS. Hexavalent chromium is cytotoxic and genotoxic to the North Atlantic right whale (Eubalaena glacialis) lung and testes fibroblasts. Mutat Research, 2008, 650, 30-38. [15] De Flora, S; Wetterhahn, KE. Mechanisms of chromium metabolism and genotoxicity. Life Chem Rep, 1989, 7, 169-244. [16] Tsuneta, Y.; Ohsaki, Y. Chromium content of lungs of chromate workers with lung cancer. Thorax, 1980, 35, 294-297. [17] Mancuso, TF. Chromium as an industrial carcinogen: Part II. Chromium in human tissues. Am J Ind Med, 1997, 2, 140-147. [18] Eastmond,DA; Macgregor, JT. Trivalent chromium: assessing the genotoxic risk of essential trace element and widely used human and animal nutritional supplement. Crit Rev Toxicol. 2008, 38, 173-190. [19] Newman, SJ; Smith, SA. Marine mammal neoplasia: A review. Vet Pathol. 2006, 43, 865-880. [20] Martineau, D; Lemberger, K. Cancer in wildlife, a case study: Beluga from the St. Lawrence Estuary, Quebec, Canada. Environ Health Perspect, 2002, 110, 285-292. [21] Al-Hamood, MH; Elbetieha, A. Sexual maturation and fertility of male and female mice exposed prenatally and postnatally to trivalent and hexavalent chromium compounds. Reprod Fertil Dev. 1998, 10, 179– 183. [22] Bataineh, H; Al-Hamood, MH. Effect of long-term ingestion of chromium compounds on aggression, sex behavior and fertility in adult male rat. Drug Chem Toxicol. 1997, 20, 133–149. [23] Chowdhury, AR; Mitra, C. Spermatogenic and steroidogenic impairment after chromium treatment in rats. Indian J Exp Biol. 1995, 33, 480–484. [24] Witmer, CM; Harris, R. Oral bioavailability of chromium from a specific site. Environ Health Perspect. 1991, 92, 105–110. [25] Witmer, CM; Park, HS. Mutagenicity and disposition of chromium. Sci Total Environ. 1989, 86, 131–148. [26] Rawson, J; Anderson, HF. Anthracosis in the atlantic bottlenose dolphin (Tursiops truncates). Mar Mamm Sci, 1991, 7, 413-416.

In: Aquatic Animals Editor: David L. Eder

ISBN: 978-1-61470-123-1 © 2012 Nova Science Publishers, Inc.

Chapter 4

HYALELLA GENUS: ANTHROPOGENIC THREATS TO A BIOINDICATOR Guendalina Turcato Oliveira*, Felipe Amorim Fernandes and Bibiana Kaiser Dutra Departamento de Ciências Morfofisiológicas - Laboratório de Fisiologia da Conservação - Programa de Pós-Graduação em Zoologia - Faculdade de Biociências da Pontifícia Universidade Católica do Rio Grande do Sul

ABSTRACT In this paper, we review the current knowledge about the usage of Hyalella genus as a bioindicator using reproductive and biochemical markers to assess damage, as well as threats (pesticides, heavy metals and endocrine disruptors) that could lead the significant alteration of the physiological pattern of these animals. Biochemical markers have the advantage of demonstrating a rapid response from animal to stress, while the reproductive show a high biological and ecological significance, because they demonstrate long-term effects that may lead to population extinction, and indicate the degree of sensitivity. Hyalella are common genus in the Americas, with 51 described species found in a variety of freshwater habitats, often cling to the vegetation, swim in the water, or burrow in the sediment where its constitute *

Bolsista de Produtividade do CNPq Email: [email protected]

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important links in the food web, serving to transfer energy from basic recourse (detritus and algae) to higher-level consumers. The structure of natural communities, the richness and diversity of species, and the interactions among them have profound implications for ecosystem functioning, modulating the fluxes of energy and materials within and among environments. The human activities have directly or indirectly added novel stressors to natural aquatic ecosystems which have the potential for affecting the structure and function of natural communities and among this stressor we can list the pesticides, the heavy metals, and home and industrial sewage that have been reported to cause dramatic changes in freshwater environments. Recently, a great deal of attention has been devoted to the use of physiological and energetic processes of non-target organisms as sensitive indicators of toxic stress. We discuss the current information on these crustaceans amphipods as models for ecotoxicology and describe areas of future research.

INTRODUCTION Members of the genus Hyalella are common in the Nearctic and Neotropical regions, with 45 described species (González and Watling 2001). They are found in a variety of freshwater habitats, such as permanent reservoirs, lakes, impoundments, and streams, and often cling to the vegetation, swim in the water, or burrow in the sediment, where they are important members of the benthic fauna for their role in trophic wiles (Kruschwitz 1978; Wellborn 1995; Grosso and Peralta 1999). Several kinds of food habits (herbivores, detritivores and planktotrophic) are present in the genus Hyalella (González et al. 2006). In the other hand, they are also important as food for other organisms, such as birds, fishes and macroinvertebrates, allowing the transfer of energy from plants to higher trophic levels (Muskó 1993; González et al. 2006). Dutra et al. (2007), studying the sympatric amphipods Hyalella pleoacuta and H. castroi in the natural environment, observed similar values (0.18±0.02 to 3.28±0.39 mmol/g) of glycogen levels for females and males of H. castroi, suggesting that the natural diet may have a high protein and low carbohydrate content. In H. pleoacuta, the levels of glycogen ranged from 1.81±0.22 to 8.76±3.16 mmol/g, suggesting a high-carbohydrate and lowprotein natural diet. This difference between these species can be explained by the behavior of H. castroi in exploiting the sediment predominantly,

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101

where it finds more organic matter of animal origin; whereas H. pleoacuta lives more in the water column, where more organic matter of plant origin is available (Dutra et al., 2007). The precise feeding habits of H. pleoacuta and H. castroi are unknown. Dutra et al. (2008a), working with H. curvispina, suggested that the natural diet of H. curvispina may have a high protein and low carbohydrate content. Other amphipods, such as Hyalella azteca, are detritivorous and herbivorous (Wen, 1992). Casset and Momo (2001) studying H. curvispina in a river of the Argentina showed that this amphipod is a herbivorous, where the principal food is phitobenthos and eventual food is sediment. Although, work developed by Dutra et al. (2011a) showed that H. curvispina are omnivorous and detritivorous, feeding algae and bacteria associated with sediments and macrophytes, and detritus present in sediment. Although the nutritional ecology of amphipods and other mesograzers is poorly understood, some feed on a range of plant, animal and detrital foods, with a few species showing a strong preference for particular prey species or groups (Nelson, 1979; Cruz-Rivera and Hay, 2000; Poltermann, 2001). Hargrave (1970) reported that H. azteca is an omnivorous deposit feeder, primarily feeding on algae and bacteria associated with the sediments and aquatic macrophytes. It has been recorded eating dead animal and plant matter (Cooper, 1965). Byrén et al. (2002) showed that for the amphipods Monoporeia affinis and Pontoporeia femorata, sedimented phytoplankton and organic detritus are their main food sources, but bacteria, meiofauna and temporary meiofauna are also included in the diet. Dutra et al. (2007) showed that the energy reserves of these hyalellids seem to be used in two different ways: (a) the adults use them for their own metabolic needs in response to the simultaneously acting environmental factors such as temperature, food availability and its composition, feeding rhythmus and others; or (b) the reserves are transferred to behavioral reproductive and to the offspring through eggs and are used by the young animals in their development. Reproductive events are important in the life cycles of these animals, leading to high energy expenditures and a close correlation with their lipoperoxidation levels. Dutra et al. (2007) showed in these amphipods (H. pleoacuta and H. castroi) that levels of lipoperoxidation may be related to reproductive behavior, motor and feeding activity and variations of the photoperiod. Environmental conditions (e.g., trophic conditions) and reproduction are supposed to be the main processes influencing the seasonal patterns of variation in biochemical composition.

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Recently, a great deal of attention has been devoted to the use of physiological and energetic processes of non-target organisms (e.g., feeding parameters, growth, respiration, reproduction, and energy-allocation mechanisms) as sensitive indicators of toxic stress from exposure to metals (Drobne and Hopkin, 1994; Khalil et al., 1995) and chemicals, including pesticides (Mohamed et al., 1992; Van Brummelen et al., 1996; Dutra et al., 2008b). The ecological relevance of these parameters is clear, because shortterm exposures can have long-term effects on the life cycles of non-target organisms, even though some compounds do not persist long in the soil (Ribeiro et al., 2001). Disturbance of the homeostasis of an organism leads to compensatory, adaptive, and finally pathological processes, which are mostly energydemanding. Therefore, the metabolic rate of an organism must increase under toxic stress (Calow and Sibly, 1990). Because the energy resources of organisms are limited, the additional metabolic costs result in a reallocation of energy resources, and can only be met at the expense of other energydemanding processes or by increased energy intake (Beyers et al., 1999).

OBJECTIVE The aim of the present review was discuss the current information of these crustaceans amphipods as modes for ecotoxicology and their physiological response to pesticide and heavy metals.

Hyalella Genus and Chemical Stressors According to Cabrera et al. (2008) agricultural production, especially on a commercial scale, has always been associated with the use of pesticides in order to prevent and combat pests that cause crop losses. The main commercially products used are synthetic organic compounds with high biological activity among these stand out, insecticides, fungicides and herbicides which are generally toxic and may be carcinogenic and cause mutations. According to Marchetti and Luchini (2004) the fate of pesticides is determined by the retention, transport and processing of these chemicals released into the environment. These processes and indirectly the transformation of these chemicals are mainly influenced by the phenomenon

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known as sorption. Sorption can influence quite significantly the leaching, a major transport process in which pesticides reach groundwater resources affecting their quality. The duration of the effect of the pesticide and its permanence in the environment set the persistence of this compound, which is also influenced by its chemical structure (Cardoso et al., 1992), and abiotic conditions and biotic environment (Silva et al. 1999). The disappearance of pesticides in soil can occur through different processes such as volatilization, chemical degradation (hydrolysis), or photo bleaching. However, in many circumstances, the loss of biological activity of a wide variety of pesticides occurs through the activity of the microbial community (Cardoso et al., 1992, Alexander, 1961). Traditionally, insecticides target nervous system functions that are common to many species, including humans. The use of neurotoxic insecticides has led to indiscriminate killing of beneficial insects and provided a serious risk to other animals and to humans through exposure to the environment (Le Couteur et al., 1999). Thus, there is a huge need to develop pesticides that pose a higher selectivity to the target and thus a reduced risk to non target species. Among biochemical markers, the inhibition of the activity of cholinesterases has been used as a diagnostic tool for organophosphate and carbamate insecticides (Ibrahim et al., 1998; Kristoff et al., 2006; Sturm et al., 1999). According Pérez et al. (2007) the structure of natural communities, the richness and diversity of species, and the interactions among them have profound implications for ecosystem functioning, modulating the fluxes of energy and materials within and among environments. The same author reported that the human activities have directly or indirectly added novel stressors to natural aquatic ecosystems which have the potential for affecting the structure and function of natural communities and among this stressor we can list the pesticides, heavy metals and urban contamination that have been reported to cause dramatic changes in freshwater environments (Pérez et al. 2007). Knight et al. (2009) were study the sediment from three Coldwater River, Mississippi backwaters was examined using 28 day Hyalella azteca bioassays and chemical analyses for 33 pesticides, seven metals and seven pesticides mixtures. Hydrologic connectivity between the main river channel and backwater varied widely among the three sites. Mortality occurred in the most highly connected backwater while growth impairment occurred in the other two. Precopulatory guarding behavior was not as sensitive as growth.

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Fourteen contaminants (seven metals, seven pesticides) were detected in sediments. Survival was associated with the organochlorine insecticide heptachlor.

Biomarkers as Means of Assessing the Toxicity Smutná (2007) reported that the use of biomarkers has been gaining recognition as means of assessing the toxicity of a medium (water or sediment) to the biota and biomarkers, which are biochemical or physiological indicators of either exposure to, or environmental contaminants at the suborganism or organism (Gillis, 1992), have a number of advantages over other methods of toxicity analysis. The same author said that like more conventional endpoints of laboratory bioassays, biomarkers provide measurement of toxicity and bioavalibility, but also can be used in the field to determine the effect of toxicants in natural environments. Smutná (2007) suggest that biomarkers can provide a link between exposure and ecologically relevant effects at a community or population level and molecular and biochemical biomarkers respond quickly to change in contaminant exposure, whereas a longer period may be required before change is apparent at population or community level (Gillis, 1992). Smutná (2007) described that one of highly important toxicity mechanisms of many xenobiotics is oxidative stress which includes pathological processes related to overproduction of reactive oxygen species (ROS) in tissues. Various modes of action can be involved in this process (Klaunig, 1998) and parameters of oxidative stress were shown to be inducible and possible to measure in invertebrates exposed to environmental stressors (Barata et al., 2005; Misra et al., 2002). According Domingues et al. (2009) since the introduction of biochemical biomarkers in Ecotoxicology, issues concerning the interpretation and validity of their various response signals have been raised. Many biomarker responses have a transient temporal feature, depending on the toxicity of the chemical, the nature of the toxic exposure, and various other factors. Hyne and Maher (2003) indicated that duration of the exposure, the rate of recovery, and the differences between pulsed and continuous exposure represented the key factors that must be understood to correctly evaluate the environmental impact of a chemical using biomarkers. According Hyne and Maher (2003) the application of biochemical measurements that can be used as individual biomarkers of impaired

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biological function in invertebrates is reviewed to evaluate whether biochemical biomarkers of aquatic invertebrates can predict changes in natural populations. Biomarkers that measure toxic effects at the molecular level have been shown to provide rapid quantitative predictions of a toxic effect upon individuals in laboratory studies (Hyne and Maher, 2003). Such biomarkers should not be used as a replacement for conventional aquatic monitoring techniques, but should be applied as supplementary approaches for demonstrating links between sublethal biochemical and adverse effects in natural populations in field studies (Hyne and Maher, 2003). The research challenge for using biomarker measurements in aquatic invertebrates is to predict effects at the population level from effects at the individual level measured upon individuals collected in the field (Hyne and Maher, 2003). According Naylor et al. (1989) and Roddie et al. (1996) the decrease in feeding rate cause by a different toxicant can be a sensitive physiological indicator of toxic stress in both freshwater and marine species, because it could be related to reductions in an organism‘s energy assimilation, which, in turn, could lead to a reduction in resource allocation to growth, reproduction, and survival, and finally translate into effects at the population level (Maltby and Naylor, 1990; Maltby, 1994; Maltby et al., 2001; Irving et al., 2003). Calow and Sibly (1990) suggested that this commonly used index of environmental stress could lead to differences in the intrinsic rate of population growth, depending on whether the reduction in reproduction was due to reduced food intake or to increased metabolic cost. Vandenbergh et al. (2003) studying the effects of the synthetic estrogen 17a-ethinylestradiol on sexual development of the freshwater amphipod Hyalella azteca exposed in a multigeneration experiment to 17aethinylestradiol concentrations ranging from 0.1 to 10 mg/L and the development of both external and internal sexual characteristics verified that in the second-generation male H. azteca exposed from gametogenesis until adulthood to 0.1 and 0.32 mg 17a-ethinylestradiol/L developed significantly smaller second gnathopods, the sex ratio of the populations exposed to 17aethinylestradiol for more than two generations tended, although not statistically significantly, to be in favor of females. The authors also described histological aberrations of the reproductive tract, i.e., indications of hermaphroditism, disturbed maturation of the germ cells, and disturbed spermatogenesis, of post-F1-generation males were observed in all 17aethinylestradiol exposures. Dutra et al. (submitted) studying the effects of acephate (organophosphate) on metabolism, lipid peroxidation, activity of the enzymes

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Na+/K+ATPase, catalase and superoxide dismutase, as well as on reproductive parameters of Hyalella castroi verified a decrease in the number of breeding pairs, absence of ovigerous females, decrease in all metabolic parameters and activity of Na+/K+ ATPase when animals were exposed to acephate. The levels of lipid peroxidation, activity of catalase and SOD showed a significantly increase compared to controls groups. The authors suggest that the results showed a strong effect of this organophosphate in this organism, which may lead to significant changes in trophic structure of liminic environments because these amphipods are important link in food chain. Dutra et al. (2008b, 2009 and 2011b) work with amphipod H. curvispina, H. pleoacuta and Hyalella castroi showed significantly mortality when animals were exposed two conc 0.36, 0.52, 1.08 and 2.16 mg/L) which was one of the most sensitive responses known. Dutra et al. (2011b) revealed that Roundup (glyphosate formulation) induces significant reduction in the levels of glycogen, proteins, lipids, triglycerides, cholesterol and activity of the enzymes Na+/K+ATPase of the Hyalella castroi showing this form a potentially toxic effect for the specie at very low concentrations that are similar to those found in water bodies. Dutra et al (2009) studying the the potential effects of carbofuran on the biochemical composition, levels of lipoperoxidation, Na+/K+ATPase activity, and reproductive behaviors (number of reproductive pairs, ovigerous females, and number of eggs) in the amphipod Hyalella castroi verfied Carbofuran induced significant decreases in biochemical reserves, a significant increase in lipoperoxidation levels, and a decrease in Na+/K+ATPase activity in both males and females. The authors suggest that studies of all the biochemical parameters seem to be quite promising, in order to assess and predict the effects of toxicants on non-target organism, as well as, reproductive behaviors may provide sensitive criteria for assessing ecotoxicological effects. H. castroi lives among rooted aquatic macrophytes, and we suggest that it is a sensitive species that could be used in monitoring studies. Giusto et al. (2010) evaluated the effect of copper on energy metabolism, lipid peroxidation levels, growth and concentration of the toxin in Hyalella curvispina and found that protein levels were not altered, the levels of glycogen, lipids, triglycerides and cholesterol were significantly decreased and levels of lipid peroxidation showed a significant increase in individuals exposed to copper, so the authors concluded that the results indicate that the

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given parameters can provide a criterion sensitive for the evaluation of ecotoxicological effects of Cu2+ on a species benthic native to the Pampas in Argentina . According Eisenhauer et al. (1999) toxicity tests are most frequently conducted on organisms derived from laboratory cultures that have been reared for many generations following standardized protocols. The restricted density of laboratory populations and the possibility of occasional population crashes increase the probability that random genetic drift and inbreeding may decrease genetic variation of the culture below that typically found in natural populations of the species. The same authors recommendate that for monitoring the genetic structure of laboratory populations is further supported by the results of studies that indicate, for some species, differential tolerance of genotypes to an environmental stressor because in the situation where certain genotypes are more resistant to a particular stressor, using a culture that is dominated by the resistant genotype may underestimate the toxic effect of a contaminant. Similarly, the inclusion of a greater number of sensitive genotypes in a test may overestimate toxicity (Eisenhauer et al., 1999).

CONCLUSION This group presented a high importance for use with pattern in toxicological test because these amphipods are important links in food chain in these habitats presented several kinds of food habits (herbivores, detritivores and planktotrophic), and they are found in a variety of freshwater habitats, such as permanent reservoirs, lakes, impoundments, and streams, and often cling to the vegetation, swim in the water, or burrow in the sediment, where they are important members of the benthic fauna. Furthermore, the results presented here in different studies showed that the amphipods, principally the specimens of the genus Hyalella are suitable organisms for use in toxicity tests, and showed that they are sensitive species that could be useful in monitoring studies for the water column and sediment.

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Dutra, BK; Fernandes, FA.; Laufer, AL; Oliveira, GT. Carbofuran-induced alterations in biochemical composition, lipoperoxidation, and Na+/K+ATPase activity of Hyalella castroi in bioassays. Comp. Biochem. Phys. C, 2009 149, 640–646. Dutra, BK; Fernandes, FA; Failace, DM. Oliveira, G.T. Effect of roundup® (glyphosate formulation) in the energy metabolism and reproductive traits of Hyalella castroi (Crustacea, Amphipoda, Dogielinotidae). Ecotox., 2011 20, 255-263. Dutra, BK; Fernandes, FA; Oliveira, GT. Carbofuran-induced alterations in biochemical composition, lipoperoxidation, and Na+/K+ATPase activity of Hyalella pleoacuta and Hyalella curvispina in bioassays. Comp. Biochem. Physiol. C, 2008b 147, 179–188. Dutra, BK; Santos, RB; Bueno, AAP; Oliveira, GT. Seasonal variations in the biochemical composition and lipoperoxidation of Hyalella curvispina (Crustacea, Amphipoda). Comp. Biochem. Physiol. A, 2008a 151, 322328. Dutra, BK; Santos, RB; Castiglioni, DS; Bueno, AAP; Bond-Buckup, G; Oliveira, GT. Seasonal variations of the energy metabolism of two sympatric species of Hyalella (Crustacea, Amphipoda, Dogielinotidae) in the southern Brazilian highlands. Comp. Biochem. Physiol. A , 2007 148, 239–247. Eisenhauer, JB; Brown, K; Sullivan, M; Lydy, J. Response of Genotypes of Hyalella azteca to Zinc Toxicity Bull. Environ. Contam. Toxicol., 1999 63, 125-132. Gillis, PL; Diener, LC. Cadmium-induced production of a metallothioneinlike protein in Tubifex tubifex (Oligochaeta) and Chironomus riparius (Diptera): Correlation with reproduction and growth. Environ. Toxicol. Chem., 1992 21(9), 1836-1844. Giusto, A; Dutra, BK; Oliveira, GT; Ferrari, L. Evaluación de efecto del cobre sobre el metabolismo energético en Hyalella curvispina. In: III Congreso Argentino de la Sociedad de Toxicología y Química Ambiental SETAC, III Congreso Argentino de la Sociedad de Toxicología y Química Ambiental SETAC, 2010 1, 65-65. Gonzáles, ER; Watling, L. Neartic and Neotropical Hyalella (Crustacea: Amphipoda: Hyalellidae). Thesis, School of Marine Science at University of Maine, USA, 2001, pp. 475. González, ER; Bond-Buckup, G; Araujo, PB. Two new species of Hyalella from southern Brazil (Amphipoda: Hyalellidae) with a taxonomic key. J. Crust. Biol., 2006 26(3),355-365.

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Grosso LE; Peralta, M. Anfípodos de água dulce sudamericanos. Revisión Del Género Hyalella Smith. I, Acta Zool. Lilloana, 1999 45, 79–98. Hargrave, BT. An Energy Budget for a Deposit-Feeding Amphipod. Limnol. Ocean., 1970 16, 99-103. Hyne, RV; Maher, WA. Invertebrate biomarkers: links to toxicosis that predict population decline, Ecotoxicol. Environ. Saf., 2003 54, 366–374. Ibrahim, H; Kheir, R; Helmi, S; Lewis, J; Crane, M. Effects of organophosphorus, carbamate, pyrethroid and organochlorine pesticides and a heavy metal on survival and cholinesterase activity of Chironomus riparius Meigen. Bull. Environ. Contam. Toxicol., 1998 60, 448-455. Irving, EC; Baird, DJ; Culp, JM. Ecotoxicological responses of the mayfly Baetis tricauadatus to dietary and waterborne cadmium: implications for toxicity testing. Environ. Toxicol. Chem., 2003 2,,1058–1064. Khalil, MA; Donker, MH; Straalen, NMV. Long-term and short-term changes in the energy budget of Porcellio scaber Latreille (Crustacea) exposed to cadmium polluted food. Eur. J. Soil Biol., 1995 31, 163–172. Klaunig, JE; Xu, Y; Isenberg, JS; Bachowski, S; Kolaja, KL; Jiang, A; Stevenson, DE; Walborg ,EF. "The role of oxidative stress in chemical carcinogenesis." Environ. Hlth. Persp., 1998 106, 289-295. Kristoff, G; Guerrero, NV; De D‘Angelo, AMP; Cochon, AC. Inhibition of cholinesterase activity by azinphos-methyl in two freshwater invertebrates: Biomphalaria glabrata and Lumbriculus variegatus, Toxicol. 2006 222, 185–194. Kruschwitz, LG. Environmental factors controlling reproduction of the Amphipod Hyalella azteca. Proc. Okla. Acad. Sci., 1978 58, 16–21. Le Couteur, DG; Mclean, AJ; Taylor, MC; Woodham, BL; Board, PG. Pesticides And Parkinson's Disease. Biom. Phamacot., 999 53, 122-130. Maltby, L. Stress, shredders and streams: using Gammarus energetics to assess water quality. In: Sutcliffe, DW. Water quality and stress indicator in marine and freshwater ecosystems. Linking levels organizations individuals, populations and communities. Windermere, Freshwater Biological Association, 1994, pp 12. Maltby, L; Kedwards, TJ; Forbes, VE; Grasman, K; Kammenga, JE; Munns, WRJ; Ringwood, AH; Weis, JS; Wood, SN. Linking individual-level response and population-level consequences. In: Baird DJ, Burton GA Jr. Ecological variability: separating natural and anthropogenic causes of ecosystem impairment. Pensacola, SETAC, 2001, pp 55.

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Maltby, L; Naylor, C. Preliminary observations on the ecological relevance of the Gammarus ‗scope of growth‘ assay: Effect of zinc on reproduction. Funct. Ecol., 1990 4, 393–397. Marchetti, ME; Luchini, LC. Pesticidas: R. Ecotoxicol. E Meio Ambiente, 2004 14, 61- 72. Misra, RB; Babu, GS; Ray, RS; Hans, RK. Tubifex: A Sensitive Model for UV-B-Induced Phototoxicity. Ecotoxicol. Environ. Saf., 2002 52, 288295. Mohamed, AI. Effects of pesticides on the survival, growth and oxygen consumption of Hemilepistus reaumuri (Audouin & Savigny 1826) (Isopoda Oniscidea). Trop. Zool., 1999 25, 145–153. Muskó, B. The life history of Corophium curvispinum G. O. Sars (Crustacea, Amphipoda) living on macrophytes in Lake Balaton. Hydrob. 1993 243, 197–202. Naylor, C; Malby, L; Callow, P. Scope for growth in Gammarus pulex, a freshwater detritivore. Hydrobiology, 1989 188, 517–523. Nelson, MK.; Brunson, EL. Postembryonic Growth And Development Of Hyalella Azteca In Laboratory Cultures And Contaminated Sediments. Chemosph., 1995 31, 3129-3140. Pérez, GL; Torremorell, A; Mugni, H; Rodríguez, P; Solange, M; Do Nascimento, VM; Allende, L; Bustingorry, J; Escaray, R; Ferraro, M; Izaguirre, I; Pizarro, H; Bonetto, C; Morris, DP; Zagarese, H. Effects of the herbicide roundup on freshwater microbial communities: a mesocosm study. Ecol. Appl., 2007 17, 2310–2322. Poltermann, M. Arctic sea ice as feeding ground for amphipods—food sources and strategies. Polar Biol., 2001 24, 89–96. Ribeiro, S; Sousa, JP; Nogueira AJA; Soares, AMVM. Effect of endosulfan and parathion on energy reserves and physiological parameters of the terrestrial Isopod Porcellio dilatatu. Ecotoxicol. Environ. Saf., 2001 49, 131–138. Roddie, BD; Redshaw, CJ; Nixon, S. Sublethal biological effects of monitoring using the common mussel, Mytilus edulis: Comparison of laboratory and in situ effects of an industrial effluent discharge. In: Tapp JF, Hunt SM, Wharfe JR. Toxic impacts of wastes on the aquatic environment. Cambridge: Royal Society of Chemistry, 1996, pp 12. Scott, S; Knight, R; Lizotte, E; Douglas, FS. Hyalella azteca (Saussure) Responses to Coldwater River Backwater Sediments in Mississippi, USA. Bull. Environ. Contam. Toxicol., 2009 83(4), 493-496.

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Silva, JM; Cunha, HL; Lopes, KC; Novato-Silva, E. Familiar Agriculture: Production Process And Health Conditions, São Paulo: Anais Do Xv Congresso Mundial Sobre Segurança E Saúde No Trabalho, 1999, pp. 40. Smutná, M. Experimental models of chronic ecotoxicity with aquatic invertebrates. Dissertation thesis in Environmental Chemistry Masaryk University, Faculty of Science, Czech Republic, RECETOX - Research Centre for Environmental Chemistry and Ecotoxicology Brno, 2007, pp112. Sturm A; Hansen, PD. Altered cholinesterase and monooxygenase levels in Daphnia magna and Chironomus riparius exposed to environmental pollutants, Ecotoxicol. Environ. Saf., 1999 42, 9–15. Van Brummelen, C; Van Gestel CAM; Verweij, RA. Long-term toxicity of polycyclic aromatic hydrocarbons for the terrestrial isopods. Oniscus asellus and Porcellio scaber. Environ. Toxicol. Chem., 1996 15, 1199– 1210. Vandenbergh, GF; Adriaens, D; Verslycke, T; Janssen, CR. Effects of 17αethinylestradiol on sexual development of the amphipod Hyalella azteca, Ecotox. Environ. Safe., 2003 54, 216–222. Wellborn, GA. Determinants of reproductive success in freshwater amphipod species that experience different mortality regimes, Anim. Behav,. 1995 50, 353–363. Wen, YH. Life history and production of Hyalella azteca (Crustacea, Amphipoda) in a hypereutrophic prairie pond in Southern Alberta. Can. J. Zool., 1992 70, 1417–1424.

In: Aquatic Animals Editor: David L. Eder

ISBN: 978-1-61470-123-1 © 2012 Nova Science Publishers, Inc.

Chapter 5

THE AROMATIC HYDROCARBON RECEPTOR MEDIATED CYTOCHROME P450 1A INDUCTION IN AQUATIC ANIMALS: BIOMONITORING OF ORGANIC POLLUTION IN AN AQUATIC ENVIRONMENT S. Arun* Alagappa University, Karaikudi-630 003, Tamil nNadu- India

ABSTRACT There is a serious concern about the disposal of hazardous substances in to the aquatic medium. Toxic contaminants such as Polycyclic Aromatic Hydrocarbons (PAHs) and Polychlorinated biphenyls Biphenyls (PCBs) are ubiquitous global contaminants that affect both the quality of water and the hygiene of aquatic organisms. Toxicity mediated through these contaminants occurs through the cytoplasmic protein called as Aromatic Hydrocarbon Receptor (AhR) which is bound to two other proteins: HSP90 and XAP2. Upon binding with the contaminant AhR translocates to the nucleus, where they dissociate from HSP90 and XAP2, and forms a heterodimer complex with bHLH-PAS and the AHR nuclear translocator (ARNT). Together, *

e-mail:[email protected].

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S. Arun they bind to Xenobiotic (contaminants / Aromatic hydrocarbons) Responsive Elements (XRE) to regulate the expression/ induction of Cytochrome P4501A (CYP1A). Thus, the induction of CYP1A through the activation of AhR has been used as a biomarker for exposure to organic pollutants in various aquatic organisms ranging from invertebrates to vertebrates. CYP1A activity has been most extensively analyzed using the ethoxyresorufin O-deethylase (EROD) or aryl hydrocarbon hydroxylase (AHH) assays. In this review article the author describes the molecular mechanism of AhR mediated CYP1A induction in aquatic organisms and how it could be applied as a suitable biomarker for the early detection of organic pollution in an aquatic environment.

1. INTRODUCTION The tremendous increase in industrialization and urbanization seriously affected the aquatic environment by way of discharging large numbers of chemicals and wastes into the aquatic medium. Toxic contaminants affect the aquatic organisms in several ways, which include changes in biochemical levels of organisms and, the simultaneously deteriorating deterioration of their health and the reducing reduction of the fitness of these aquatic organisms. Failures of in the toxicity defense mechanisms and repair systems in aquatic organisms leads to the death of these individuals, which further affects the populations, and even communities. Study of the consequences of the toxic contaminants on aquatic organisms and aquatic systems is known as aquatic toxicology. In particular, aquatic toxicology deals with the interactions between environmental toxic contaminants and biota that focuses on the adverse effects at different levels of biological organization such as molecular, cellular, tissue and organ levels of aquatic organisms [1]. For the past few decades, large numbers of works were carried out to reveal the mechanisms of toxic contaminants in aquatic organisms at the cellular level. In particular, ecotoxicological research on selected contaminants requires an interdisciplinary effort, considering biochemical, molecular, toxicological, physiological and ecological processes. Toxic contaminants interact with the biological systems of aquatic organisms through a variety of mechanisms that give rise to physical, chemical, and biochemical modifications of biological macromolecules or changes in gene expression. The most important mechanism is that in which the initial step occurs through toxic contaminants interaction with intracellular receptor proteins which act as ligand-dependent

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transcription factors. Nowadays, many of these receptor proteins appear to have evolved as biological sensors to detect the presence of toxic contaminants in the aquatic organisms or in the aquatic medium. It was found that these receptors often regulate the expression of genes encoding xenobiotic metabolising enzymes. This includes phase I biotransformation enzymes such as cytochromes P450 and microsomal flavin monooxygenases. The induction of such genes can usually be considered an adaptive response. Many research works related to the induction of these enzymes in aquatic organisms in response to toxic contaminants revealed that cytochrome P450 enzymes could be used as biomarkers to identify the toxic effects in aquatic organisms. The roles of cytochrome P450 in activating toxic contaminants have been defined mostly in mammalian systems. However, fish and invertebrates possess microsomal enzymes, including cytochrome P450 and monooxygenase enzymes and the basic biochemistry of the microsomal cytochrome system and its functions in aquatic species have been explained in detail in earlier reviews [2-5]. Main The main role of cytochrome P-450 systems in both fish and mammals is the inducibility of P450 enzymes by toxic chemicals which are structurally related compounds and aid in the biotransformation of chemicals. Toxic contaminants mediated mediating induction of cytochrome P450 has been demonstrated in numerous studies. Table 1 indicates the variety of compounds that can act as inducers in aquatic organisms. Table 1. Variety of Compounds that can Act as Inducers in Aquatic Organisms Toxic chemicals / Contaminants Benzo(a)pyrene Dibenzanthracene Methylcholanthrene 3,3 ',4,4'-Tetrachlorobiphenyl 2,3,7,8-Tetrachlorodibenzodioxin B-Naphthoflavone Hexachlorobiphenyl Polybrominated biphenyl Pentachloro biphenyl Polychlorinated naphthalene

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It has been identified that a ligand-activated nuclear receptor known as Aromatic Hydrocarbon Receptor (AHR) plays a role in the regulation and induction of cytochrome P450 (CYP) genes [6]. Particularly in aquatic organisms, the Cytochrome P4501A (CYP1A) induction occurs through contaminants mediated AhR pathway. In brief, the toxic contaminant binds with AhR complex (AhR complex comprises of HSP90 and XAP2) and translocates into the nucleus. Then, AhR dissociates from HSP90 and XAP2, and forms a heterodimer complex with bHLH-PAS and the AHR nuclear translocator (ARNT). Together, they bind to Xenobiotic Responsive Elements (XRE) to regulate the expression/ induction of Cytochrome P4501A (CYP1A). Thus, the induction of CYP1A through the activation of AhR has been well studied in aquatic organisms and used as a biomarker to organic pollutants in various aquatic organisms ranging from invertebrates to vertebrates. CYP1A activity has been most extensively analysed using the ethoxyresorufin O-deethylase (EROD) or aryl hydrocarbon hydroxylase (AHH) assays. In this review article, the molecular mechanism of AhR mediated CYP1A induction in aquatic organisms is described and the way it could be applied as a suitable biomarker for the early detection of organic pollution in aquatic environment is explained.

2. CYTOCHROME P450 – ORIGIN AND EVOLUTION Cytochrome P450 is a heme protein and it is generally known as being in the supergene family of enzymes that defends against xenobiotic compounds such as environmental contaminants. Cytochrome P450s have been found in all classes of organisms except anaerobic bacteria. It is originated from an ancestral gene that existed approximately 3.5 billion years ago [7]. Up tTo date, there were athave been at least 500 P450 genes and 70 families had have been classified on the basis of amino acid sequence similarity. It is believed that multiple genes can be expressed frequently in all organisms. Due to genetic multiplicity and lack of preservation within the P450 gene family in organisms, the naming and assignment of individual genes into families and subfamilies is based on levels of amino acid sequence identity as determined by the P450 Nomenclature Committee. In brief, the CYP450 gene/ cDNA sequence was mentioned beginning with the italicized root ―CYP‖ and followed by an Arabic number representing the family, a capital letter representing the sub family, then by another Arabic number for

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individual gene (eg., CYP1A1, CYP6B1). Similarly, the same nomenclature is used for CYP protein/mRNA but the names are not italicized (CYP1A1, CYP6B1). Cytochrome P450 originated 3.5 billion years ago and the CYP super family is believed to undergone several rounds of modification by gene duplication. Mitochondrial P450 (CYP11) was formed 1.5 billion years ago from this super family that are was primarily involved in the metabolism of cholesterol. About 900 million years ago, another expansion of the gene family occurred and this family is believed to be involved in steroids synthesis (CYP19,CYP21 and CYP27). Later, drug metabolizing enzymes have been expanded (CYP1 and CYP2). About 400 million years ago a dramatic expansion of xenobiotic metabolizing gene families were formed and it has been hypothesized that this recent formation has been driven by at least two major events. The first was the emergence of aquatic vertebrates onto land which leads led to the introduction of toxic allelochemicals in to their diets. The second was hydrocarbon hydrocarbon-based products released by the combustion of plant materials [8]. Thus the CYP mediated monooxygenase gene family (CYP1A) evolved from the CYP superfamily and is involves involved in the metabolism of PAH‘s.

Based on Hollebone and Brownlee [14]. Figure 1. Mechanism of xenobiotic compound metabolism by Cytochrome P450.

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3. MECHANISM OF CYTOCHROMEP 450 About 60 years ago, Klingenberg [9] and Garfinkel [10] revealed Cytochrome P450 as a pigment having a novel carbon monoxide binding spectrum in liver microsomes of rats and pigs. Then Omura and Sato [11] summarized an evident mechanism and the history of P450. Cytochromes P450 which comprises a large superfamily of heam-containing enzymes. These enzymes are of particular importance when studying the detoxification and metabolism of toxic contaminants in aquatic animals. The cytochrome P450 serves as both the oxygen and substrate-binding locus for the monooxygenase reaction and undergoes cyclic oxidation / reduction of the heme iron during catalysis. It consists of a single polypeptide chain with ironprotoporphyrin IX loosely bound by hydrophobic forces, and electrostatic and covalent bonds. The cytochrome P450 monooxygenase function in association with several microsomal proteins, located in the endoplasmic reticulum together with the P450 isoenzymes that, transfer electrons to reduce cytochrome P450 during its catalytic cycle. These components include NADPH-cytochrome c reductase, cytochrome b5, and NADHcytochrome c reductase. Electrons from NADPH or NADH are transferred through these compounds and cytochrome P450 (CYP 450) inserts one atom of oxygen into the toxic contaminants and reduces the second oxygen atom to form water. There have been two major views on the function of CYP 450 enzymes i.e., it is involved in the metabolism of endogenous compounds (steroid hormones, vitamins and bile acids) and exogenous compounds such as toxic contaminants and drugs [12]. The mechanism of xenobiotic metabolism by heme protein (CYP 450) was clearly explicated by Hollebone [13] and Hollebone and Brownlee [14] (Figure 1). In this model, the lipid bound protein is an oligomeric globin, capable of both rotation and translation within the lipid. As some sites of the globin protein would face the cytosol, oxygen molecules can be captured from the aqueous medium of the cell. By applying antibonding electron density to this trapped oxygen at the heme site, the protein is capable of heterolysis such as an oxygen molecule to a hydroxide ion, which is released back to the cytosol, and a positive oxygen atomic ion, which remains bound to the active site by back back-bonding to the iron (II) state of the heme. This activated site is protected by a tyrosine cover. After rotation of the globin and opening of the lipophilic tyrosine, this hydrophobic, activated site can encounter and adsorb hydrophobic toxic contaminants that dissolved in the lipid bilayer. Then, the oxygen positive ion oxidizes the contaminants. After oxidation, the

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hydrophilic hydroxylated contaminant‘s metabolite could force rotation of the protein and permit release of the metabolites into lumen for conjugation reaction and detoxification.

4. CYTOCHROME P450 IN AQUATIC ORGANISMS The Cytochrome was first discovered in the liver by Klingenberg [9] and Carfinkel Garfinkel [10] as a carbon monoxide binding pigment and named P450 from the position at 450 nm of the absorption peak of the carbon monoxide complex of the reduced pigment [11]. The metabolism of xenobiotic compounds to various oxidative derivatives was demonstrated in mammals as early as the 1940‘s. Early studies of detoxification systems in aquatic organisms seemed to indicate that they were incapable of microsomal oxidations [15] but in the late 1960‘s, a few studies demonstrated that aquatic organisms possessed oxidative enzymes which were effective in the transformation of xenobiotic compounds [16,17]. Subsequently, the existence of cytochrome P450 dependent monooxygenase activities was well documented in fish [18-22]. Complementary functions of these enzymes in metabolism or conversion of toxic xenobiotic compounds such as polycyclic aromatic hydrocarbons, polychlorinated biphenyls pesticides and other toxic compounds was were also studied in fishes [23-29]. Despite considerable number of works being carried out on CYP450 and monooxygenase enzymes such as CYP1A, fewer studies were conducted on aquatic invertebrates in related relation to vertebrate organisms. An attempt was made on sponges for detecting CYP related biotransformation enzymes. But the activity of CYP monooxygenase enzymes was not detected in sponges like Tethya aurantium, Verongia aerophoba, Pellina semitubulosa, Tethya limski and Suberitus domuncula [30]. Similarly, CYP mediated metabolism of PAH was not detected in coelenterates [31]. In 1984, it was found that, CYP content was not present in detectable level in nematodes, Haemonchus contortus [32]. But later in 1997 Kotz and co-workers [33] found CYP450 as well as detoxification mechanisms in nematodes [34]. In the earlier portion of the 1980‘s, the CYP450 induction towards xenobiotics has had been extensively studied in annelida species [35]. Elmarnlouk et al. [36] observed the presence of CYP450 in crustacean species. In the mollusk, Livingstone et al. [37] assessed the CYP450 mediated xenobiotic detoxification.

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5. CATALYTIC CYCLE OF CYP450 AND MONOOXYGENASE REACTION Four intermediates of the CYP450 catalytic cycle are well characterized and reviewed by Mansuy [38]. In its resting state, CYP450 generally exists as a mixture of a hexacoordinate low spin Fe(III) state with a water molecule trans to its endogenous cysteinate liagand (1 in Figure 2) and a pentacoordinate high spin Fe (III) state with the csysteinate only (2 in Figure 2). Binding of substrates in a protein site shifts the equilibrium between the two Fe(III) states towards the pentacoordinate complex. The third intermediate complex CYP450 Fe(II) is pentacoordinate high spin complex (3 in Figure 2). It can bind with various ligands including O2. As mentioned earlier CYP450 reactive species are mainly involved in the transfer of an atom of oxygen into substrates. This complex is the fourth intermediates of CYP450 and is known as the high-valent iron –oxo complex (4 in Figure 2). Mansuy [38] also assessed another intermediate, of CYP450 catalytic cycle, –

ferric peroxo complex CYP450 Fe(III)-O-O (3a in Figure 2). This complex is believed to be involved in the P450 catalysed oxidative cleavage of C-C and C=N bonds. As mentioned in the earlier paragraph (mechanism of CYP450), the CYP450 dependent monooxygenase reaction includes the transfer of one atom of oxygen into substrates. A Few few examples of monooxygenase reactions are given below for the interest of readers. The electrophilic P450 iron-oxo intermediate inserts its oxygen into inerted C-H bonds of alkanes. CYP450 also is able to catalyse the hydroxylation of N-H bonds of amines and amidines. Similarly, it also transfers its oxygen atom to electron rich centers of substrates such as aromatic rings and carbon-carbon double bonds which leads to generation of reactive metabolites such as epoxides and arene oxides. These reactive metabolites may react with protein and glutathione and causes oxidativce damages to biological organisms.

6. CYP450 MEDIATED XENOBIOTIC (BENZO(A)PYRENE) METABOLISM IN AQUATIC ORGANISMS Metabolites of Polycyclic Aromatic Hydrocarbons are well known carcinogenic chemicals and are among the various PAH‘s the model compound benzo(a)pyrene (BaP) that has received the most attention in

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aquatic organisms, with a primary focus on hepatic metabolism and adduct formation. Metabolites of BaP have been shown experimentally to be carcinogenic to aquatic organisms following both waterborne and oral exposures. So, in this review article we paid more attention to the impact of B(a)P on aquatic organisms and their metabolizing ability. In addition, for the convenience of readers, the generation of B(a)P metabolites in vertebrate and invertebrate organisms was described with two models such as fish and crustaceans.

Figure 2. Catalytic cycle of Cytochrome P450.

Several studies have investigated CYP mediated BaP (polycyclic aromatic hydrocarbons) metabolism by microsomes of both aquatic vertebrate and invertebrate species (Figure 3).

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Figure 3. Pathway of Benzo(a)pyrene metabolism.

In fish, benzo ring metabolites of BaP were predominantly observed. Several positions of oxygenation were observed and quinones were often major metabolites [39]. In aquatic organisms, CYP1A from several species has high activity with BaP and preferentially metabolizes in the benzo ring at the 7,8- and 9,10- positions, whereas other cytochromes P450, such as the CYP2 and CYP3 family isoforms usually have lower activity and metabolize BaP preferentially at the 3 position. Kleinow et al. [39] also detected B(a)P metabolites such as 3-OH-BaP, 9-OH-BaP, BaP-7,8-diol, BaP-9-sulfate and BaP-9-glucuronide in cat fish. However, earlier studies generally indicated poor metabolism of BaP in the invertebrate species, compared with fish species [40]. There are many works to support the above evidences. For example, microsomes from the freshwater crab, Eriocheir japonicus, had BaP-3-hydroxylase (AHH) activities ranging from undetectable to 20 pmol mim–1 mg–1 protein [41]. Similarly, freshwater crustaceans had detectable levels of ethoxyresorufin –o- deethylase activity (CYP1A) and involves involved low metabolism of PAH’s [82]. Thus, the B(a)P metabolizing ability is not stronger in invertebrate species compared to fishes. In general, in invertebrates, the CYP mediated B(a)P metabolites production are lower or not found in traceable levels. When, the hepatic microsomes of freshwater crustaceans M. malcolmsonii, was incubated with B(a)P, very low levels of 3-OH B(a)P and 7,8 dihydro diol metabolites were detected (unpublished results from author’s research findings; Figure-4). A wide range of PAH metabolizing capacity has been observed in marine invertebrates such as Mollusca (Hydrobia totteni, Ilyanassa obsoleta, Yoldia limatula, and Gemma gemma), Annelida (Nereis succinea, Pectinaria

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gouldii, Haploscolopolous sp., and Capitella sp.1) and Arthropoda (Edotea triloba, and Gammarus mucronatus) [40]. Despite the ability of many invertebrates to metabolize PAHs, the general lack of responsiveness of most invertebrates to antibodies raised against cytochrome P450 (CYP) 1A in fish [42] and the apparent lack of an Ah receptor in many members of this group [43] raise questions concerning the role of CYP1A and its regulation in invertebrate species. The details about the presence of Ah Receptor and its role of CYP1A induction in vertebrates and invertebrates are explained in next paragraph.

Unpublished results of Author‘s research work. Figure 4. Detection of Benzo(a)pyrene metabolites in crustacean species.

7. AROMATIC HYDROCARBON RECEPTOR (AHR) AND ITS MECHANISM The Aromatic Hydrocarbon Receptor (AhR) is believed to mediate the CYP metabolism of PAH‘s and dioxin. In 1976, Poland and co-workers [44] located the AhR in mouse livers, where it was found specifically and with high affinity to bind to the radiolabelled analogue of TCDD, a toxic compound. At that time, numbers of research works have beend started to carried carriy out tests on the mechanisms of AhR action and its connections

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with the toxic responses of xenobiotic compounds such as PCB‘s and PAH‘s. Before going getting to know about the mechanism of AhR, let us know learn more about the structure of AhR. The structure of AhR required many years of study to be identified and confirmed, due to the instability of the receptor protein and low levels of expression in biological organisms. Biochemical purification of the AhR protein succeeded only after development of a photoaffinity-labelled ligand [45,46]. The sequence information of purified AhR protein was then used to clone the cDNA [47,48]. Later on, AhR cDNAs were cloned from several animals, including mammals, birds, fish and some invertebrates [49]. The AhR cDNA cloning revealed that it belongs to the basic helix-loop-helix/PER-ARNT-SIM (bHLH/PAS) superfamily, which is a complex consisting of three members of proteins. These proteins play a main role in the adaptation of environmental change. The first member, ARNT is a dimerization partner of the AHR. PER is a second member and is known as a product of the Drosophila Period gene. The third one is SIM, a product of the Drosophila Single-minded locus. SIM shares a highly conserved PAS domain [50,51]. The AHR has a conserved aminoterminus (N terminus). The bHLH domain located nearest the Nterminus binds DNA and helps to dimerise with the ARNT. The PAS domain affords specificity for dimerization and also contains most of the ligandbinding domain (LBD) [52,53]. Apart from this, both the bHLH and PAS domains are responsible for interaction with the 90-kDa heat shock protein (hsp90) [54]. Since, N-terminal end contains nuclear localization (NLS) and the nuclear export signals (NES), the AhR shuttles between the cytoplasm and the nucleus of the cell [55,56]. The C-terminal end of the AHR contains a potent transactivation domain composed of several interacting subdomains and an important one is a glutamine-rich (Q-rich) region (Figure-5 - based on the review of Hahn [43]).

Figure 5. Structure of Ah Receptor protein.

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As mentioned earlier, the AhR protein is present in the cytoplasm as a complex with two molecules of hsp90 and some other protein. In addition to this, there are also suggestions that some auxiliary co-chaperones, such as p60 and hsp70 are bound with AhR protein. Ligands for the AhR ie dioxin or PAH enter into the cells by simple diffusion and bind to the PAS B domain. The AhR complex disassociates from hsp70 protein and then translocates into the nucleus and heterodimerizes with the ARNT. This AHR-ARNT heterodimer generates the complex as in the high affinity DNA-binding form. This complex binds to its specific DNA recognition site (DRE; Dioxin Responsive Element), located upstream of the target genes, such as CYP1A1 [57,58,59]. Thus, upon binding with DRE, it regulates the expression of CYP1A (Figure 6)

Figure 6. AhR mediated Induction mechanism of CYP1A.

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(TCCD: 3,4,7,8-tetrachlorodibenzodioxin; AhR: Aromatic hydrocarbon receptor; HSP90:Heat shock protein 90; ARNT:Aromatic hydrocarbon nuclear transferaselocator; XRE: Xenobiotic response element).

8. AHR IN AQUATIC ORGANISMS In 1985, many works have been carried out to find out AhR protein in non non-mammalian species and but failed to find out thea presence of AhR due to some problems in laboratory techniques [57,58]. They have used [3H] TCDD and velocity sedimentation on sucrose gradients for to identifying identify the AhR protein. However, later Heilmann et al. [60] investigated the existence of AhR in rainbow trout fish cells. Then, several works have beenwere carried out on the presence of AhR in fish (Table 2). Different kinds of assays are used for the detection of AhR and photoaffinity labeling was identified as a good technique for the detection of AhR in fish. It is now investigated found that, the AhR of fish differs from mammals by way of having two AhR genes namely AhR1 and AhR2. These two AhR have been found in zebrafish [61,62], medaka [63], and two species of pufferfish [64] rubripes and Tetraodon fluiatilis) [65]. Existence and xenobiotic binding capacity of invertebrate‘s AhR remain stillstill remain unclear. Few A few years back, AHR homologs have been identified in molluscsmollusks, including the soft shell clam Mya arenaria, the zebra mussel Dreissena polymorpha and the blue mussel Mytilus edulis [66,67]. It is very interesting to note that, like AHRs from D. melanogaster and C. elegans, the mollusc mollusk AHR homologs have bHLH and PAS domains and they interact with mammalian AHRE sequences, and fail to bind to [3H]TCDD or [3H]BNF. Since, the AhR ligand is unable to bind with invertebrate‘s AhR, xenobiotic mediated CYP1A induction in aquatic invertebrates is still unanswerable. In the year of 1999, CYP1A protein has beenwas identified in marine mussels by Livingstone and his co-workers. Microsomes of mussels collected from the polluted area exhibit induced CYP1A. When the microsomes were cross cross-reacted with polyclonal anti-perch and monoclonal anti anti-fish CYP1A antibodies, immunopositive protein was clearly detected in Mytilus edulis [68] (Figure 7). So, extensive work is warranted in connection with invertebrate‘s AhR protein and its role on CYP1A induction.

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Figure 7. Detection of CYP1A immunopositive protein in mussel, Mytilus edulis.

Table 2. List of marine/freshwater organisms that posses Ah Receptor protein S. No 1 2 3 4 5

Species Killifish (Fundulus heteroclitus) Rainbow trout (Oncorhynchus mykiss) Zebrafish (Danio rerio) Medaka (Oryzias latipes) Zebrafish (Danio rerio)

References [69] [70] [71] [72] [61]

6 7 8

Medaka (Oryzias latipes) Pufferfish (Fugu Abe) Soft-shell clam (Mya arenaria)

[63] [64] [66]

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Table 3. CYP1A enzyme in vertebrates and invertebrates as a biomarker to pollutants Site

Organisms

Contaminants

Reference

---

Biomarker (CYP1A) AHH

Avalon Penisula, Canada River Rhone, France Lake Saimaa, Finland

Brown trout Nase

PCB

AHH

[75]

Rainbow trout

AHH

[76]

Perch

Pulp and paper mill effluent Oil Spill

Gulf of Bothnia, Vaasa, Finland Thames estuary, UK Erka Oil Spill Spain

EROD

[77]

Eel

PAH

EROD

[78]

Asterias rubens Clams and Crabs

PAH PAH

CYP1A EROD

[79] [80]

[74]

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9. CYP1A AS A BIOMARKER TO ENVIRONMENTAL POLLUTION The catalytic activity of Cytochrome P4501A (ethoxyresorufin-odeethylase) is considered as a suitable biochemical marker for the detection of PAH contamination. Payne et al. [73] was first introduced the CYP450 protein as a biomarker to organic pollution. After that, many works have been emerged in relation to CYP4501A and environmental biomarkers. Particularly, various numbers of works have had been carried out on the ethoxyresorufin-o-deethylase activity of fish and this enzyme was used as a successful biomarker for marine pollution. For the interest of readers, a few examples are wereare furnished in Table 3. When compared to vertebrates, invertebrate species exhibit low rates of metabolism for the elimination of contaminants. So, readily metabolizable contaminants such as PAH‘s tend to bioaccumulate to the highest tissue concentration at the bottom of food chains in invertebrates. In vertebrates, it has been extensively studied and reported that the induction of cytochrome P4501A (ethoxyresorufin –O- deethylase) catalytic activity is successfully involved in the metabolism of polycyclic aromatic hydrocarbons. The AhR functions as a ligand activated transcription factor that is involved in the expression of cytochrome P4501A and its associated monooxygenase activity. There is no detailed information available on invertebrate‘s AhR functional relationship with hydrocarbons and CYP1A inducibility [49]. Therefore, limited number of works is are available concerning the presence of CYP1A (EROD activity) in invertebrates, especially on crustaceans and mollusks. Arun et al. [82] reported the presence of CYP1A catalytic activity (EROD) in the larvae, sub sub-adult and adult stages of crustacea (M. malcolmsonii). In addition, the response of EROD activity in M malcolmsonii has been assessed, following experimental exposure to sub-lethal concentration of oil effluent. The accumulated levels of hydrocarbon contents, in the tissues of crustacean, gradually increased and attained higher values at day 30. Similarly, the EROD activity was gradually elevated and a maximum value of 284 ± 6.8 pmol hr-1 mg-1 protein was observed at day 30. This enhanced activity showed a three fold hike when compared to control (p